d^k United States 

Environmental Protection 
L^l M % Agency 


TD 756 
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.052 
2002 
Copy 2 


FT MEADE 
GenCol1 


Anaerobic Compost 
Constructed Wetlands System 
(CWS) Technology 


Innovative Technology 
Evaluation Report 






















EPA/540/R-02/506 
December 2002 



Anaerobic Compost 
Constructed Wetlands System 
(CWS) Technology 


Innovative Technology Evaluation Report 


National Risk Management Research Laboratory 
Office of Research and Development 
U.S. Environmental Protection Agency 
Cincinnati, Ohio 45268 



Recycled/Recyclable 

Printed with vegetable-based ink on 
paper that contains a minimum of 
50% post-consumer fiber content 
processed chlorine free. 



Notice 


The information in this document has been funded by the U. S. Environmental Protection Agency (EPA) under Contract No. 
68-C5-0037 to Tetra Tech EM Inc. It has been subjected to the Agency's peer and administrative reviews and has been 
approved for publication as an EPA document. Mention of trade names or commercial products does not constitute an 
endorsement or recommendation for use. 


/ ^ 75 ^ 

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Foreword 


The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation's land, air, and water 
resources. Under a mandate of national environmental laws, the Agency strives to formulate and implement actions leading 
to a compatible balance between human activities and the ability of natural systems to support and nurture life. To meet this 
mandate, EPA's research program is providing data and technical support for solving environmental problems today and 
building a science knowledge base necessary to manage our ecological resources wisely, understand how pollutants affect our 
health, and prevent or reduce environmental risks in the future. 

The National Risk Management Research Laboratory is the Agency's center for investigation of technological and 
management approaches for reducing risks from threats to human health and the environment. The focus of the Laboratory's 
research program is on methods for the prevention and control of pollution to air, land, water and subsurface resources; 
protection of water quality in public water systems; remediation of contaminated sites and ground water; and prevention and 
control of indoor air pollution. The goal of this research effort is to catalyze development and implementation of innovative, 
cost-effective environmental technologies; develop scientific and engineering information needed by EPA to support 
regulatory and policy decisions; and provide technical support and information transfer to ensure effective implementation of 
environmental regulations and strategies. 

This publication has been produced as part of the Laboratory's strategic long-term research plan. It is published and made 
available by EPA's Office of Research and Development to assist the user community and to link researchers with their clients. 


Hugh W. McKinnon, Director 

National Risk Management Research Laboratory 


ill 



Abstract 


As part of the Superfund Innovative Technology Evaluation (SITE) Program, the U.S. Environmental Protection Agency (EPA) 
evaluated constructed wetlands systems (CWS) for removing high concentrations of zinc from mine drainage at the Burleigh 
Tunnel in Silver Plume, Colorado. 

Exploration geologists have known for many years that metals, most commonly copper, iron, manganese, uranium, and 
zinc, frequently accumulate in swamps and bogs located in mineralized areas. This understanding forms the basis for the 
design of CWS—essentially excavated pits filled with organic matter—that have been developed and constructed over the 
past 15 years to treat drainage from abandoned coal mines in the eastern United States. Mine drainage is routed through 
the organic material, where metals are removed through a combination of physical, chemical, and biological processes. 

In fall 1994, anaerobic compost wetlands in both upflow and downflow configurations were constructed adjacent to and 
received drainage from the Burleigh Tunnel, which forms part of the Clear Creek/Central City Superfund site. The 
systems were operated over a 3-year period. The effectiveness of treatment by the CWS was evaluated by comparing the 
concentration of zinc and other metals from corresponding influent and effluent analyses. By far the dominant toxic metal 
present in the drainage was zinc. The upflow CWS removed an average of 93 percent of the zinc during the first year of 
operation, and 49 and 43 percent during the second and third years. The downflow CWS removed an average of 77 
percent of zinc during the first year and 70 percent during the second year. (Flow was discontinued to the downflow 
system in the third year.) 


IV 



Contents 


List of Figures and Tables.viii 

Acronyms, Abbreviations, and Symbols.ix 

Conversion Factors.xi 

Acknowledgments.xii 

Executive Summary. 1 

1 Introduction.5 

1.1 Brief Description of the SITE Program and Reports.5 

1.2 Purpose of the Innovative Technology Evaluation Report.6 

1.3 Technology Description.6 

1.3.1 Treatment Technology.8 

1.3.2 System Components and Function.8 

1.3.3 Key Features of the CWS Technology.9 

1.4 Key Contacts. 11 

2 Technology Application Analysis. 12 

2.1 Applicable Wastes.12 

2.2 Factors Affecting Performance.12 

2.2.1 Mine Drainage Characteristics.12 

2.2.2 Operating Parameters.13 

2.2.3 Compost Performance. 13 

2.3 Site Characteristics. 13 

2.3.1 Support Systems.13 

2.3.2 Site Area, Preparation, and Access. 15 

2.3.3 Climate.15 

2.3.4 Utilities.15 

2.3.5 Services and Supplies. 15 

2.4 Availability, Adaptability, and Transportability of Equipment. 15 

2.5 Material Handling Requirements. 16 

2.6 Personnel Requirements. 16 

2.7 Potential Community Exposures. 16 

2.8 Evaluation of Technology Against RI/FS Criteria. 16 

2.9 Potential Regulatory Requirements. 18 


v 


































Contents (continued) 


2.9.1 Comprehensive Environmental Response, Compensation, and Liability Act. 18 

2.9.2 Resource Conservation and Recovery Act. 18 

2.9.3 Clean Water Act. 19 

2.9.4 Occupational Safety and Health Act. 19 

2.10Limitations of the Technology. 19 

3 Treatment Effectiveness.22 

3.1 Background.22 

3.2 Review of SITE Demonstration.22 

3.2.1 Treatability Study.22 

3.2.2 Technology Demonstration.23 

3.2.3 Operational and Sampling Problems and Variations from the Work Plan.23 

3.2.4 Site Demobilization.24 

3.3 Demonstration Methodology.24 

3.3.1 Testing Approach.25 

3.3.2 Sampling, Analysis, and Measurement Procedures.25 

3.4 Site Demonstration Results.27 

3.4.1 Burleigh Mine Drainage Chemistry.27 

3.4.2 Downflow CWS.27 

3.4.3 Upflow CWS.36 

3.4.4 Clear Creek.40 

3.4.5 Toxicity Testing Results.40 

3.4.6 Microbial Toxicity Testing.42 

3.5 Attainment of Demonstration Objectives.43 

3.6 Design Effectiveness.44 

3.6.1 Downflow Cell.44 

3.6.2 Upflow Cell.45 

4 Data Quality Review.46 

4.1 Zinc Data Quality Review.46 

4.1.1 Quality Assurance Results for Field Sampling Activities.46 

4.1.2 Quality Assurance Results for Sample Analysis.47 

4.2 Acute Toxicity Data Quality Review. 48 

4.2.1 Analytical Quality Assurance.48 

4.3 Noncritical Parameters Data Quality Review.50 

5 Economic Analysis.. 


VI 





































Contents (continued) 


5.1 Basis of Economic Analysis.52 

5.2 Cost Categories.53 

5.2.1 Site Preparation Costs.53 

5.2.2 Permitting and Regulatory Requirements.53 

5.2.3 Capital Equipment.53 

5.2.4 Startup.55 

5.2.5 Labor.55 

5.2.6 Consumables and Supplies.55 

5.2.7 Utilities.55 

5.2.8 Residual Waste Shipping and Handling.55 

5.2.9 Analytical Services.55 

5.2.10 Maintenance and Modifications.55 

5.2.11 Demobilization.56 

6 Technology Status.57 

7 References.58 

Appendix 

A Analytical Results Summary Tables 
B Case Study 


vii 


















1 

2 

3 

4 

5 

6 

7 

8 

9 

10 

11 

1 

2 

3 

4 

5 

6 

7 

8 

9 


.. 3 

. 5 

17 

21 

23 

24 

25 

27 

29 

32 

33 

11 

13 

20 

26 

31 

35 

35 

36 

47 


Figures 


Site Location. 

Schematic Cross-Section of an Anaerobic CWS Upflow Cell 

Flow Rates Measured for Effluent Cells. 

CWS Zinc Concentration by Month. 

CWS Cadmium Concentration by Month. 

CWS Lead Removal by Month. 

CWS Manganese Removal by Month. 

Sulfate-Reducing Bacteria, Downflow CWS Substrate. 

Monthly Zinc Loading, Downflow CWS. 

Sulfate-Reducing Bacteria, Upflow CWS Substrate. 

Monthly Zinc Loading, Upflow CWS. 


Tables 

Evaluation of CWS Treatment Versus RI/FS Criteria. 

Treatment Standards and Influent concentrations for the CWS SITE Demonstration 

Summary of Standard Methods and Procedures. 

Average Downflow CWS Substrate Results. 

Average Upflow CWS Substrate Results. 

Clear Creek Upstream. 

Clear Creek Downstream. 

CWS Demonstration Toxicity (LC ?0 ) Results. 

CWS Costs for Different Treatment Flow Rates. 


viii 























Acronyms, Abbreviations, and Symbols 


°c 

Degrees Celsius 

°F 

Degrees Fahrenheit 

%C 

Percent completeness 

%R 

Percent recovery 

AA 

Atomic absorption 

ARAR 

Applicable or relevant and appropriate requirement 

ASTM 

American Society for Testing and Materials 

AVS 

Acid volatile sulfide 

BOD 

Biochemical oxygen demand 

CDPHE 

Colorado Department of Public Health and Environment 

CDM 

Camp, Dresser, & McKee, Inc. 

CFU 

Colony forming units 

CERCLA 

Comprehensive Environmental Response, Compensation, and Liability Act 

CFR 

Code of Federal Regulations 

CWS 

Constructed wetlands system 

DQO 

Eh 

Data quality objective 

Oxidation reduction potential 

EPA 

U.S. Environmental Protection Agency 

FS 

Feasibility study 

gpm 

h 2 s 

HDPE 

Gallons per minute 

Hydrogen sulfide 

High-density polyethylene 

HSWA 

Hazardous and Solid Waste Amendments of 1984 

ICP 

Inductively coupled plasma 

ITER 

Innovative technology evaluation report 

lc 50 

MCAWW 

Lethal concentration for 50 percent of the test organisms 

Methods for Chemical Analysis of Water and Wastes 

MCL 

Maximum contaminant level 

Pg 

Micrograms 


IX 



Acronyms, Abbreviations, and Symbols (continued) 


pS 

Microsiemens 

mg/kg 

mg/L 

MS 

Milligrams per kilogram 

Milligrams per liter 

Matrix spike 

NCP 

National Oil and Hazardous Substances Pollution Contingency Plan 

NIST 

National Institute of Standards and Technology 

NPDES 

National Pollutant Discharge Elimination System 

NRMRL 

National Risk Management Research Laboratory 

O&M 

Operation and maintenance 

ORD 

Office of Research and Development 

ORP 

Oxidation/reduction potential 

OSHA 

Occupational Safety and Health Administration 

OSWER 

Office of Solid Waste and Emergency Response 

PPE 

Personal protective equipment 

ppm 

PRC 

Parts per million 

PRC Environmental Management, Inc. 

PVC 

Polyvinyl chloride 

QAPP 

QA/QC 

RCRA 

Quality assurance project plan 

Quality assurance/quality control 

Resource Conservation and Recovery Act 

RI 

Remedial investigation 

RPD 

Relative percent difference 

SARA 

Superfund Amendments and Reauthorization Act 

SITE 

Superfund Innovative Technology Evaluation 

SDWA 

Safe Drinking Water Act 

SOP 

Standard operating procedure 

SRM 

Standard reference material 

SWDA 

Solid Waste Disposal Act 

TCLP 

Toxicity characteristic leaching procedure 

TOC 

Total organic carbon 

TDS 

Total dissolved solids 

TSS 

Total suspended solids 

yd 3 

Cubic yards 


X 



Conversion Factors 


Length 


Area: 


Volume: 


Mass: 


Energy: 


Power: 


To Convert From 

To 

Multiply By 

inch 

centimeter 

2.54 

foot 

meter 

0.305 

mile 

kilometer 

1.61 

square foot 

square meter 

0.0929 

acre 

square meter 

4,047 

gallon 

liter 

3.78 

cubic foot 

cubic meter 

0.0283 

pound 

kilogram 

0.454 


kilowatt-hour megajoule 3.60 


kilowatt horsepower 1.34 


Temperature: 


(°Fahrenheit - 32) 


“Celsius 


0.556 



Acknowledgments 


This report was prepared under the direction of Mr. Edward Bates, the U.S. Environmental Protection Agency (EPA) Superfund 
Innovative Technology Evaluation (SITE) project manager at the National Risk Management Research Laboratory (NRMRL) 
in Cincinnati, Ohio; Ms. Dana Allen, U.S. EPA Region VIII; and Mr. James Lewis, Colorado Department of Public Health 
and Environment. This report was prepared by Mr. Gary Miller, Mr. Garry Farmer, Mr. Jon Bridges, and Ms. Shaleigh 
Whitesell of Tetra Tech EM Inc. (Tetra Tech) and Mr. Mark Kadnuck of the Colorado Department of Public Health and 
Environment (formerly of Tetra Tech). This report was typed by Ms. Robin Richey and Ms. June Diller, edited by Mr. Butch 
Fries, and reviewed by Dr. Kenneth Partymiller of Tetra Tech. 

This project consisted of a demonstration conducted under the SITE program to evaluate the anaerobic compost Con¬ 
structed Wetland System (CWS) technology developed by the Colorado Department of Public Health and Environment 
(CDPHE). The technology demonstration was conducted on mineral mine drainage at the Burleigh Tunnel in Silver 
Plume, Colorado, which is included in the Clear Creek/Central City Superfund site. Passive treatment was selected as the 
preferred treatment alternative for the Burleigh Tunnel drainage in a 1991 Record of Decision (ROD). This Innovative 
Technology Evaluation Report (ITER) interprets the data that was collected during the nearly four-year demonstration and 
discusses the potential applicability of the technology. 

The cooperation and participation of the following people are gratefully acknowledged: Mr. Vincent Gallardo, Ms. Ann 
Vega, and Dr. James Lazorchek of NRMRL; Ms. Holly Fliniau of EPA Region VIII and Mr. Rick Brown of CDPHE. 


xii 



Executive Summary 


This executive summary of the Constructed Wetlands 
System (CWS) technology demonstration discusses 
technology applications, describes system effectiveness, 
and presents an evaluation of the costs associated with the 
system and lessons learned during the field demonstration. 

Introduction 

The anaerobic compost CWS technology was evaluated 
under the Superfund Innovative Technology Evaluation 
(SITE) program. The SITE program was developed by 
the U.S. Environmental Protection Agency (EPA) in 
response to the mandate of the Superfund Amendments 
and Reauthorization Act (SARA) of 1986. The primary 
purpose of the program is to maximize the use of alternative 
treatment technologies. To this end, reliable performance 
and cost data on innovative technologies are developed 
during demonstrations where the technology is used to 
treat a specific waste. 

After the demonstration, EPA publishes an Innovative 
Technology Evaluation Report (ITER) designed to aid 
decision makers in evaluating the technology for further 
consideration as an appropriate cleanup option. This 
ITER includes a review of the technology application, an 
economic analysis of treatment costs, and the results of 
the demonstration. 

For this demonstration, wetlands were designed and 
constructed to treat mine drainage through a combination 
of physical, chemical, and biological processes. The mine 
drainage, containing primarily zinc contamination, flowed 
into the constructed wetlands where metals were removed 
by sorption, precipitation, and biological sulfate reduction. 
The demonstration included the evaluation of two CWS 
treatment cells (pits) filled with an organic-rich compost 
(96 percent) and alfalfa hay (4 percent) mixture. Both 
treatment cells were constructed adjacent to, and received 
drainage from, the Burleigh Tunnel in Silver Plume, 


Colorado. The Burleigh Tunnel is part of the Clear Creek/ 
Central City Superfund site. Passive wetlands treatment 
was identified by the Colorado Department of Public 
Health and Environment (CDPHE) as the preferred 
remedial alternative for the Burleigh Tunnel drainage. 

Each treatment cell covered 0.05 acres and differed in 
flow configuration. One cell was constructed in an up flow 
configuration, in which water entered from the base of the 
cell and was forced upward to discharge; the other was 
constructed in a downflow configuration, in which water 
entered from the top of the cell and flowed by gravity to 
discharge. The compost and hay mixture was 4 feet deep 
in both cells. Flow rates of water into and out of the cells 
were controlled by a series of v-notch weirs; each cell 
was designed to treat 7 gallons per minute (gpm). 

Technology Applications Analysis 

The primary objectives of the CWS technology 
demonstration were to (1) measure the reduction of zinc 
in Burleigh Tunnel drainage resulting from the CWS 
treatment with respect to cell configuration and seasonal 
variation (temperature); (2) assess the toxicity of the 
Burleigh Tunnel drainage; (3) characterize the toxicity 
reduction resulting from treatment of the drainage by the 
CWS; and (4) estimate toxicity reductions in the stream 
(Clear Creek) receiving the Burleigh Tunnel drainage. 

CWS treatment effectiveness was evaluated by comparing 
the concentration of zinc and other metals from 
corresponding CWS influent and effluent analyses 
(see Section 3.0). The results indicate the concentration 
of zinc in the Burleigh Tunnel drainage ranged from 50 to 
60 milligrams per liter (mg/L) during the first year of the 
demonstration. However, in May and June 1995, a great 
deal of spring snow and rain and a rapid thaw combined 
to increase the amount of runoff entering the mine 
network drained by the Burleigh Tunnel. At that time, 


1 



flow from the tunnel increased from 45 gpm to more than 
300 gpm, and zinc concentrations increased from 55 mg/ 
L (April 12, 1995) to 109 mg/L (August 8, 1995). Over 
the final 2 years of the demonstration, zinc concentrations 
in Burleigh Tunnel mine drainage were lower in the 
winter, dropped again in April or May when flow through 
the mine workings increased, and rapidly increased in 
summer, remaining high throughout the fall. During this 
period, Burleigh Tunnel mine drainage zinc concentrations 
generally remained between 45 and 84 mg/L, with 
increases to more than 100 mg/L noted during the late 
summer and fall. The Burleigh Tunnel drainage is also 
characterized by moderate pH and alkalinity and low 
concentrations of metals other than zinc. 

Downflow 

In the first year of operation, CDPHE reported the 
downflow cell developed flow problems on occasion, 
preventing treatment of the intended amount of water. 
Remedies, such as fluffing the compost, were tried and 
were somewhat successful allowing the system to flow at 
4 to 6 gpm during the first two years of operation. During 
the third year, the flow in this cell dropped to less than 
1 gpm and flow to this cell was discontinued. 
The permeability loss is believed to be related to 
precipitation of metal oxides, hydroxides, and carbonates, 
settling of fine materials in the cell, and compaction of the 
substrate material. 

In general, the downflow cell was effective in removing 
zinc during the first year of operation. Zinc removal by this 
cell ranged from 69 to 96 percent with a mean removal of 
77 percent. During the second year of operation, zinc 
removal ranged from 62 to 79 percent with a mean of 70 
percent. During the final 6 months of operation, flow 
through the downflow cell continued to decline increasing 
the residence time of the mine drainage in the cell. The 
increased residence time improved zinc removal. Zinc 
removal during this period ranged from 67 to 93 percent 
with a mean of 82 percent. 

Aqueous geochemical modeling, observations of cell 
compost, sulfate-reducing bacteria count results, and acid 
volatile sulfide data suggest that biological sulfate reduction 
is not the primary zinc removal mechanism within this cell. 
Instead, the primary metal removal mechanism is thought 
to be the precipitation of zinc oxides, hydroxides, and 
carbonates in aerobic sections of the downflow cell. 


Upflow 

During the first 6 months of operation, upflow cell effluent 
samples contained low (less than 1 mg/L) concentrations 
of zinc. However, during the later part of 1994 and into 
1995, upflow cell effluent zinc concentrations began to 
increase. The concentrations of zinc ranged from 0.13 mg/ 
L in early 1994 to 60.1 mg/L in May 1997. 

In the spring of 1995, heavy spring runoff overwhelmed 
the CWS system, channeling 20 gpm of aerobic water 
(nearly three times the design flow) through the upflow 
cell. This high runoff also apparently mobilized more zinc 
from the mine workings or mine waters and substantially 
increased the concentration of zinc in the mine drainage. 
The large flows created aerobic conditions and the 
increased zinc loading had a detrimental effect on the 
upflow cell. These new conditions apparently initiated a 
change in the cell’s microbial ecology. After the high flow 
event, the upflow cell removed only 50 to 60 percent of the 
zinc in the mine drainage. Prior to the high flow event, the 
upflow cell removed greater than 90 percent of the zinc 
contamination (year 1 mean removal was 93 percent). 

The loss of substrate hydraulic conductivity also affected 
the upflow CWS. During the demonstration, the height of 
the influent wier was periodically raised to increase the 
hydraulic pressure to maintain flow through the upflow 
CWS. The water level was raised approximately 1 foot 
over the 4-year demonstration. In 1997, this cell developed 
a visibly obvious preferential pathway in the southeast 
comer, adjacent to the bermed sidewall. This preferential 
pathway was eliminated by terminating flow to this section 
of the wetland through excavating of the wetland substrate 
to allow installation of a cap on the influent line. 

The high initial zinc removal rates in the upflow cell were 
likely the result of absorption of metals and biological 
sulfate reduction. The decline in metal removal by the 
upflow cell after the high flow event is likely related to the 
decline in sulfate reducing bacteria in this cell. There are 
several possible reasons for the decline of the sulfate- 
reducing bacteria including toxicity produced by high zinc 
concentrations for the bacteria, prolonged exposure to 
aerobic conditions allowing other wetland bacteria to 
outcompete the sulfate-reducing bacteria, or the utilization 
of all the most readily metabolized growth materials by the 
sulfate reducing bacteria leading to lower activity and 
eventually lower populations of these bacteria. Ultimately, 
the primary metal removal mechanism over the last 
several years of the demonstration was likely chemical 
precipitation. 


2 




Economic Analysis 

An economic analysis was conducted to examine 11 cost 
categories for the CWS technology. The 11 categories 
include (1) site preparation; (2) permitting and regulatory 
requirements; (3) capital equipment and construction; 
(4) startup; (5) labor; (6) consumables and supplies; 
(7) utilities; (8) residual and waste shipping and handling; 
(9) analytical services; (10) maintenance and 
modifications; and (11) demobilization. 

A number of factors affect the estimated costs of treating 
mine drainage with the CWS technology. These factors 
generally include flow rate, type and concentration of 
contaminants, water chemistry, physical site conditions, 
site location, and treatment goals. In addition, the 
characteristics of the spent compost produced by a CWS 
will affect disposal costs since the compost may require 
treatment for off-site disposal. 

Based on the criteria evaluated in the cost analysis, the 
average estimated cost for a constructed wetland at 
50 gallons per minute (gpm) over a 15-year period is 
$ 1,744,100 million or $0.0045 per gallon of water treated. 

Treatment Effectiveness 

Based on this demonstration, the following conclusions 
may be drawn about the effectiveness of the anaerobic 
compost CWS technology. 

• The upflow CWS removed an average (arithmetic 
mean) of 53 mg/L (93 percent) of zinc during the 
first year of operation. 

• Upflow cell zinc removal averaged 41 mg/L 
(49 percent) during the second year and 30 mg/L 
(43 percent) during the third year of operation. 

• During the first year of operation, the upflow cell 
effluent was not toxic to Ceriodaphnia dubia or 
Pimephales promelas. 

• The downflow CWS removed an average of 44.2 mg/ 
L (77.4 percent) of zinc during the first year and 58 
mg/L (70 percent) during the second year of 
operation. 

• The CWS is relatively easy to construct with readily 
available materials. 

In summary, results from this SITE demonstration and 
additional tests of the CWS technology suggest that the 
CWS is capable of reducing the toxicity of contaminated 
mine drainage by removing metals such as zinc, cadmium, 
iron, lead, nickel, and silver. 


However, the results of this demonstration also clearly 
show that an anaerobic compost CWS using sulfate 
reduction may have difficulty in recovering from upset 
conditions such as the high flow event that occurred 
during this demonstration. 

In addition, application of this technology to mine drainage 
containing high concentrations of iron may require 
pretreatment to remove the iron. If not removed, the iron 
could precipitate in the wetland and could lead to loss of 
wetland permeability. 

Lessons Learned 

The following items highlight lessons learned during the 
CDPHE constructed wetlands demonstration. The list is 
partitioned among five categories of considerations (or 
concerns): theory, design, construction, operation and 
maintenance, and analytical. 

Theory 

• An upflow CWS using biological sulfate reduction is 
capable of reducing the concentration of several 
metals including zinc, cadmium, nickel, lead, iron, and 
silver. The extent of metal reduction depends on the 
concentration of the metal and sulfate in the mine 
drainage, and the performance of the CWS. 

• The primary metal removal process in the downflow 
CWS did not appear to be biological sulfate reduction. 
Zinc in the demonstration CWS downflow cell 
appeared to be primarily removed by chemical 
precipitation. Generally, zinc removal by the 
demonstration downflow cell ranged between 70 and 
80 percent. However, the accumulation of zinc 
carbonate in the cell compost may have attributed to 
a loss of cell permeability during the demonstration. 

Design 

• A hydraulic residence time of 50 hours (estimated) 
provided good metal removal in the upflow cell 
during the first 8 months of the demonstration. 
However, the decline in metal removal after this 
initial period and inability to re-establish the sulfate- 
reducing bacteria in the upflow cell suggest this 
residence time may be a lower limit for mine drainages 
containing high metal concentrations. 

• Hydraulically, the upflow cell performed well with 
4 feet of compost. However, some short circuiting 
was observed after 3 years of operation. 

• The mixture of fresh compost (96 percent) and 
hay (4 percent) used as a substrate during the 
demonstration was a superior environment for sulfate- 
reducing bacteria. However, the compost contains 
high levels of ammonia that is readily leached during 


3 



wetland startup, resulting in elevated levels of 
ammonia in the discharge. The addition of wood 
products to the substrate can reduce the amount of 
ammonia generated. Land treatment has been used 
at some sites to dispose of wetland startup discharge. 

• Each wetland cell should have an easily adjustable 
influent conveyance with the capability of bypassing 
200 to 300 percent of typical peak flows. 

Construction 

• Bermed sidewalls lined with high-density polyethylene 
(HDPE) is a suitable construction technique for cold 
region applications. However, the use of a geonet 
on the wetland surface to allow animals and people 
to walk on the wetland is not recommended. The 
geonet did not allow additional compost or hay to be 
added to the wetland. In addition, the use of geofabric 
to separate the piping networks from the compost is 
not recommended. 

• Effluent collection pipes (polyvinyl chloride [PVC]) 
should be larger than 1 inch in diameter to prevent 
clogging from precipitated material. In addition, the 
effluent collection structure should include cleanouts 
that allow precipitated material to be periodically 
removed without driving the precipitate back into the 
wetland compost. 

• Lining a downflow cell with HDPE above the level 
of the ponded water allows this water to short circuit 
the wetland compost. Short circuits are most 
noticeable during the winter when the compost 
becomes frozen and contracts from the liner, creating 
a gap between the compost and liner. 

Operation and Maintenance 

• Constructed wetlands can require frequent inspections 
to ensure that proper flows are maintained within the 
treatment cells. However, properly designed and 
constructed influent distribution and effluent collection 
networks may reduce inspection frequency. 

• Treatment system downtime with CWS treatment is 
not high. Effluent piping networks should be cleaned 
out periodically (once or twice a year was appropriate 
for the Burleigh Tunnel CWS). The frequency of 
compost removal and replacement will depend on 
contaminant loading, metal removal efficiencies, and 
the desired performance level of the CWS. Compost 
removal and replacement frequency for the 
demonstration CWS upflow cell is estimated to be 
once every 4 to 5 years. 

• Straw bales covered with insulated construction 
blankets (used to cure concrete in cold weather) are 
an effective insulator for an upflow CWS during 
winter operation. However, their use requires an 
upper support structure such as a geonet. An 
equally effective insulation system could include 


6 inches of fresh compost and hay covered by 
construction blankets, although this system has not 
been tested. 

• Straw bales used for winter insulation must not be 
allowed to become saturated by water. Their 
combined weight will compress the wetland compost, 
making it impermeable. 

Analytical 

• Routine (monthly) total metals analysis in 
conjunction with quarterly dissolved metals analysis 
were useful in evaluating the performance of the 
CWS. The mine drainage and effluents were 
sampled and analyzed every 2 weeks during the first 
2 years of the demonstration; however, monthly 
sampling (conducted over the final year of the 
demonstration) is adequate to track treatment 
performance. 

• Routine aquatic toxicity testing of the mine drainage 
and CWS effluent also provides useful water quality 
information. During the CWS demonstration, these 
analyses were conducted every 3 to 4 months, but 
semi-annual analyses could also be used. 
Demonstration aquatic toxicity testing used two test 
organisms, fathead minnows (Pimephalus promelas) 
and water fleas (Ceriodaphnia dubia); however, other 
test organisms including trout fry could also be used. 

• Sulfate-reducing bacteria analyses of wetland 
compost were conducted monthly during the first 2 
years of the CWS demonstration. These analyses, 
while useful, did not show much variation until the 
high flow event, and their frequency could easily be 
reduced to every other month or even a quarterly. 
Acid volatile sulfide analysis can indicate the 
accumulation of metal sulfides within the CWS 
compost; however, the compost sample must be 
collected from the area of metal filtration. The acid 
volatile sulfide analysis procedure is not routine for 
most laboratories, and meaningful results may not be 
achievable. 

• All aqueous field analyses conducted during the 
CWS demonstration including pH, Eh (effluent), 
dissolved oxygen (influent), conductivity, and 
temperature were useful measurements. It should 
be noted that the platinum element of the Eh probe is 
prone to poisoning, requiring periodic replacement. 


4 



Section 1 
Introduction 


This section provides background information about the 
SITE program, discusses the purpose of this ITER, and 
describes the C WS technology. Key contacts for additional 
information about the SITE program, this technology, and 
the demonstration site are listed at the end of this section. 

1.1 Brief Description of the SITE 
Program and Reports 

SARA mandates that EPA select, to the maximum extent 
practicable, remedial actions at Superfund sites that create 
permanent solutions (as opposed to land-based disposal) 
for contamination that affects human health and the 
environment. In response to this mandate, the SITE 
program was established by EP A’s Office of Solid Waste 
and Emergency Response (OSWER) and Office of 
Research and Development (ORD). The SITE program 
promotes the development, demonstration, and use of 
new or innovative technologies to clean up Superfund 
sites across the country. 

The SITE program’s primary purpose is to maximize the 
use of alternatives in cleaning up hazardous waste sites by 
encouraging the development and demonstration 
of innovative treatment and monitoring technologies. It 
consists of the Demonstration Program, the Emerging 
Technology Program, the Monitoring and Measurement 
Technologies Program, and the Technology Transfer 
Program. These programs are discussed in more detail 
below. 

The objective of the Demonstration Program is to develop 
reliable performance and cost data on innovative treatment 
technologies so that potential users may assess specific 
technologies. Technologies evaluated either are currently 
or will soon be available for remediation of Superfund 
sites. SITE demonstrations are conducted at hazardous 
waste sites under conditions that closely simulate full- 
scale remediation, thus assuring the usefulness and 
reliability of information collected. Data collected are 


used to assess the performance of the technology, the 
potential need for pre- and post-treatment processing of 
wastes, potential operating problems, and approximate 
costs. The demonstrations also allow evaluation of long¬ 
term risks and operating and maintenance (O&M) costs. 

The Emerging Technology Program focuses on 
successfully proven, bench-scale technologies that are in 
an early stage of development involving pilot-scale 
or laboratory testing. Successful technologies are 
encouraged to advance to the Demonstration Program. 
The constructed wetlands is an example of a successful 
graduate of the Emerging Technology Program that was 
evaluated in the Demonstration Program. 

Existing technologies that improve field monitoring and 
site characterization are identified in the Monitoring and 
Measurement Technologies Program. New technologies 
that provide faster, more cost-effective contamination 
and site assessment data are supported by this program. 
The Monitoring and Measurement Technologies Program 
also formulates the protocols and standard operating 
procedures for demonstrating methods and equipment. 

The Technology Transfer Program disseminates technical 
information on innovative technologies in the 
Demonstration, Emerging Technology, and Monitoring 
and Measurement Technologies Programs through various 
activities. These activities increase the awareness and 
promote the use of innovative technologies for assessment 
and remediation of Superfund sites. The goal of technology 
transfer is to promote communication among remedial 
managers requiring up-to-date technical information. 

Technologies are selected for the SITE Demonstration 
Program through annual requests for proposals. ORD 
staff review the proposals, including any unsolicited 
proposals that may be submitted throughout the year, to 
determine which technologies show the most promise for 
use at Superfund sites. Technologies chosen must be at 


5 



the pilot- or full-scale stage, must be innovative, and must 
have some advantage over existing technologies. Mobile 
technologies are of particular interest. Once EPA has 
accepted a proposal, cooperative agreements between 
EPA and the technology developer establish responsibilities 
for conducting the demonstrations and evaluating the 
technology. The developer is responsible for demonstrating 
the technology at the selected site and is expected to pay 
any costs for transportation, operation, and removal of 
equipment. EPA is responsible for project planning, site 
preparation, sampling and analysis, quality assurance and 
quality control (QA/QC), and for preparing reports, 
disseminating information, and transporting and disposing 
of untreated and treated waste material. For the CWS 
evaluation, CDPHE (the lead agency of the Burleigh 
Tunnel site) identified passive wetlands treatment as the 
preferred treatment alternative with agreement by EPA 
and the division of responsibilities was essentially as 
described. 

The results of the CWS technology demonstration are 
published in two documents: the SITE technology capsule 
and the present ITER. The SITE technology capsule 
provides relevant information on the technology, 
emphasizing key features of the results of the SITE field 
demonstration. The ITER is discussed in the following 
section. Both the SITE technology capsule and the ITER 
are intended for use by remedial managers making a 
detailed evaluation of the technology for a specific site and 
waste. 

1.2 Purpose of the Innovative 

Technology Evaluation Report 

The ITER provides information on the CWS technology 
and includes a comprehensive description of the 
demonstration and its results. The ITER is intended for 
use by EPA remedial project managers, EPA on-scene 
coordinators, contractors, and other decision makers for 
implementing specific remedial actions. The ITER is 
designed to aid decision makers in evaluating specific 
technologies for further consideration as an option in a 
particular cleanup operation. This report represents a 
critical step in the development and commercialization of 
a treatment technology. To encourage the general use of 
demonstration technologies, EPA provides information 
regarding the applicability of each technology to specific 
sites and wastes. Therefore, the ITER includes information 
on cost and site-specific characteristics. It also discusses 
advantages, disadvantages, and limitations of the 
technology. Each SITE demonstration evaluates the 
performance of a technology in treating a specific waste. 


The waste characteristics at other sites may differ from 
the characteristics of the treated waste. Therefore, 
successful field demonstration of a technology at one site 
does not necessarily ensure that it will be applicable at 
other sites. Data from the field demonstration may 
require extrapolation for estimating the operating ranges 
in which the technology will perform satisfactorily. Only 
limited conclusions can be drawn from a single field 
demonstration. 

1.3 Technology Description 

The Colorado Department of Public Health and 
Environment submitted a proposal to the SITE program 
for demonstrating the anaerobic compost CWS technology. 
This technology was selected for a SITE demonstration at 
the Burleigh Tunnel in Silver Plume, Colorado. The 
demonstration was carried out under a cooperative 
agreement involving the EPA National Risk Management 
Research Laboratory (NRMRL), CDPHE, and EPA 
Region 8. 

The Burleigh Tunnel is located approximately 50 miles 
west of Denver in the Silver Plume - Georgetown mining 
district (Figure 1), within the Clear Creek/Central City 
Superfund site. The Silver Plume - Georgetown mining 
district occupies an area of about 25 square miles 
surrounding the towns of Silver Plume and Georgetown. 
The tunnel entrance is at an elevation of 9,152 feet, about 
400 feet north of Clear Creek, on the western side of the 
town of Silver Plume. The area immediately surrounding 
the tunnel entrance is littered with mill tailings and waste 
rock dumps. Dilapidated buildings and equipment 
from previous milling operations are also present. 
No mining operations are active in the immediate area. 
The water draining from the Burleigh Tunnel is of near¬ 
neutral pH (ranging from 6.9 to 7.9) and has high zinc 
concentrations (ranging from 44.8 to 109 mg/L). The 
drainage also contains moderate alkalinity and low levels 
of metals other than zinc. 

A treatability study was conducted at the Burleigh Tunnel 
between June 18, 1993 and August 12, 1993. The 
treatability study involved the construction, operation, and 
sampling of two up flow compost and hay bioreactors that 
treated mine drainage from the Burleigh Tunnel. The 
treatability study (PRC 1993) showed that low levels of 
sulfate in the mine drainage would not limit biological 
sulfate reduction, thereby permitting the removal of zinc 
and other metals by the bioreactors or the demonstration 
scale treatment cells. Construction of the CWS 
demonstration cells began in August 1993 and was 


6 







Summit ( 
County 


Park 

County 



5 Ml 0 5 Ml 10 Ml 


SCALE: 1” = 10 MILES 


Figure 1 . Site location. 


7 






























































completed in November 1993. The demonstration began 
in January 1994 and continued for a 46-month period 
through November 1997. Evaluation of the CWS 
technology is based on results of the treatability study and 
the SITE demonstration at the Burleigh Tunnel site. 

1.3.1 Treatment Technology 

There are generally three types of constructed wetlands: 
free-water surface systems, subsurface flow systems, 
and aquatic plant systems (EPA 1988). A free-water 
system typically consists of shallow basins or channels 
with slow- flowing water and plant life. A subsurface 
flow wetland consists of basins or channels filled with 
permeable substrate material; the water flows through, 
rather than over, this substrate. An aquatic plant system 
is essentially a free water surface system with deeper 
channels containing floating or suspended plants. In 
general, free-water surface and aquatic plant systems are 
aerobic wetlands that remove metals primarily by aerobic 
oxidation of iron followed by precipitation of iron hydroxides, 
that leads to the precipitation or adsorption of other 
metals. Aerobic wetlands are most successful in removing 
iron, arsenic, selenium and, to some extent, manganese 
from moderately low to neutral pH mine waters (Gusek 
and others 1994). 

Anaerobic compost wetlands are designed to treat mine 
drainage through a combination of physical, chemical, and 
biological processes. Mine drainage is directed into 
constructed wetlands that contain an organic-rich compost 
substrate. Initially, sorption to the CWS substrate is the 
primary metal removal mechanism active within the 
system. Sorption includes adsorption of metals to organic 
and inorganic wetlands materials and absorption of metals 
into wetlands microorganisms and plants. 

• Adsorption refers to the binding of positively charged 
ions to mineral surfaces by metal cations in solution. 
The sorption of inorganic ions is largely determined 
by complex chemical equilibria involving the charge 
and size of the element or complex ion, the nature of 
the sorbing material, and the pH of the aqueous 
solution. The properties of the surface that influence 
inorganic soiption include net surface charge and the 
presence, configuration, and pH dependence of 
binding sites. The structure of the solid may also 
affect adsorption reactions. 

• Absorption refers to the incorporation of ions 
or compounds into the cell structure of 
microorganisms or plants. Metals may also be 
incorporated into the structure of complex humic 
substances formed during the degradation of the 
substrate. 


After several months, the sorption capacity of the wetlands 
is exhausted and metal removal efficiencies by this 
mechanism decline. 

Once the sorption capacity of the CWS substrate is 
expended, the formation, precipitation, and filtration of 
metal sulfides become the primary metal removal 
mechanism in the CWS. The process is believed to be 
biologically mediated by sulfate-reducing bacteria present 
in anaerobic zones within the CWS. 

The bacteria oxidize organic matter provided by the 
wetland with the simultaneous reduction of sulfate to 
hydrogen sulfide. The hydrogen sulfide reacts with 
dissolved metals to produce metal sulfides. The metal 
sulfides, with low aqueous solubilities, precipitate and 
become trapped in the wetlands substrate by filtration. 
The following reactions illustrate the overall oxidation/ 
sulfate reduction reactions and subsequent formation of 
metal sulfides. 

SO; 2 + 2CH O —> HS- + 2HCO; + H + 

4 2 3 

M +2 + H 2 S or HS’ —> MS(s) + 2H* 

where: M is a metal such as zinc (Zn +2 ), iron (Fe +2 ), nickel 
(Ni +2 ), and (s) indicates a solid. 

In addition, other reactions within the wetlands may 
contribute to observed metal removal, including mineral 
precipitation and chelation (binding) to suspended organic 
material. In general, mine drainage contains low levels of 
dissolved oxygen that, when exposed to air, will take up 
oxygen and become aerobic. This process can lead to 
geochemical disequilibrium where the metal is no longer 
soluble at this concentration and may initiate metal 
precipitation. Zinc carbonate (Smithsonite) is an example 
of a mineral that may precipitate in the demonstration 
downflow CWS. In addition, the decay of wetland 
compost and biomass will produce dissolved and suspended 
organic material in the wetland pore water. These 
materials can chelate metals in solution. Although chelated 
metals may not be effectively removed (filtered) by the 
wetland, they may not be available biochemically to 
aquatic plants and organisms exposed to the effluent. 

1.3.2 System Components and Function 

Two CWS treatment cells were located adjacent to the 
Burleigh Tunnel between a compressor building and an 
old mill. Each cell covered 0.05 acre; the two cells 
differed in flow configuration. The cell nearest the mine 


8 



adit was an upflow system, in which water entered the cell 
under pressure from the bottom and flowed upward 
through the substrate material to discharge. The second 
cell was a downflow system, in which the water entered 
the cell from the top and flowed by gravity to the bottom 
for discharge. The demonstration CWS cells were highly 
engineered systems compared to many of the previously 
tested constructed wetlands, including the Big 5 wetlands 
evaluated in the Emerging Technology Program (EPA/ 
540/R-93/523). Figure 2 shows a cross-section schematic 
of the upflow CWS treatment cell. The downflow cell 
was identical except the direction of mine drainage flow 
in the compost is reversed. 

Both CWS treatment cells were installed below grade to 
reduce freezing of the cells during winter. Both had 
bermed earthen side walls. The base of each cell was 
made up of a gravel subgrade, a 16-ounce geofabric, a 
sand layer, a clay liner, and a high density polyethylene 
liner. The base was separated from the influent or 
effluent piping by a geonet. A 7-ounce geofabric separated 
the perforated PVC piping from the compost. The compost 
was held in place with a combination of 7-ounce geofabric 
and geogrid in the upflow cell. The perforated effluent 
piping was also supported by the geogrid in the upflow cell. 
Up to 6 inches of dry substrate material was located above 
the perforated piping. The geonet and the perforated 
piping ensured even distribution of the influent water into 
the treatment cells and prevented short circuiting of water 
through the cells. The influent and effluent distribution 
piping were also staggered horizontally as an additional 
precaution against short circuiting. 


• Upflow cell - 69 feet long, 25.5 feet wide, and 4 feet 
deep, with an estimated total substrate volume of 
198 cubic yards 

• Downflow cell - 62 feet long, 33 feet wide, and 4 
feet deep, with an estimated total substrate volume 
at 218 cubic yards 

Note: The dimensions listed are at the top of the cell 
wall The volumes listed take into account the sloped 
walls of the cells. 

The organic-rich compost substrate was composed of a 
mixture of 95 to 96 percent manure compost and 4 to 
5 percent hay. The compost was produced from cattle 


manure and unidentified paper products. The compost 
and hay mixture had been identified as the most effective 
medium in removing zinc from the drainage during the 
previous bench-scale test (Camp, Dresser and McKee 
1993). Wood based substrates have also been used in 
constructed wetland systems. 

The flow to the CWS cells was regulated by a series of 
concrete v-notch weirs, one for the influent and one for 
the effluent of each cell. The effluent weir controlled the 
flow and the hydraulic residence time of the mine drainage 
through both CWS cells. Each cell was designed for a 
flow of 7 gpm with a total flow capacity for the two cells 
of 14 gpm. The remaining flow from the Burleigh Tunnel 
drainage was diverted to Clear Creek (untreated) via the 
influent weir. A drainage collection structure was 
constructed within the Burleigh Tunnel to build sufficient 
hydraulic head to drive the flow through the two CWS. 

1.3.3 Key Features of the CWS 
Technology 

Certain features of the CWS technology allow it to be 
adapted to a variety of settings: 

• The hardware components (geosynthetic materials, 
PVC piping, and flow control units) of the CWS are 
readily available. 

• Compost materials can be composed of readily 
available materials. However, the actual composition 
of a substrate material for a site-specific constructed 
wetland is best detennined through pilot studies. 
Composted manure was used during this study. 

Operation and maintenance costs are low since the 
systems are generally self-contained, requiring only 
periodic changes of the compost depending on site- 
specific conditions. 

Other features that should be thoroughly evaluated before 
constructing a CWS include the following: 

• Properties of the drainage to be treated. Some 
drainages may need some type of pretreatment 
before entering the CWS. For example, drainage 
with high iron or aluminum content might prematurely 
clog the CWS if not pretreated to remove some of 
the metal. 

• Climate conditions must be evaluated to assess the 
potential for reduced efficiency of the system during 
different seasons of the year. 

• Contingencies if the system does not perform as 
expected. 


Existing construction near the Burleigh Tunnel entrance 
required that the upflow cell be 10 percent smaller by 
volume than the downflow cell. The dimensions of the 
cells are as follows: 


9 




Figure 2. Schematic cross-section of an anaerobic CWS upflow cell. 


10 


















































Proximity to a populated area—odors generally are 
associated with CWS treatment. 


The Clear Creek/Central City Superfund Site 


• Land availability near the source of the contaminated Michael Holmes - Remedial Pro J ect Manager 
water to avoid extended transport. The CWS U.S. Environmental Protection Agency 
typically requires more land than a conventional Region 8 

treatment system. Consequently, locations with 999 18th Street, Suite 300 
steep slopes and drainages would make construction o enver Colorado 80202 
more difficult and costly. Telephone: (303)312-6607 

• Cost of constructing the system if substrate and 
other materials are not readily available. 

• Possible use of concrete basins to eliminate 
replacement costs for liners. 

• Potential for vandalism of the CWS, which could 
result in increased costs. 

• Seasonal fluctuation of water flow or chemistry and 
the potential impact to the CWS. 

• Production and release of nutrients from substrate 
and stream standard requirements for discharge of 
produced nutrients 

1.4 Key Contacts 

Additional information on the CWS technology, the SITE 
program, and the demonstration site can be obtained from 
the following sources: 

The CWS Technology 

James Lewis 

Colorado Department of Public Health and Environment 

HMWMD-RP-82 

4300 Cherry Creek Drive South 

Denver, Colorado 80222-1530 

Telephone: (303)692-3390 

Fax: (303)759-5355 

The SITE Program 


Edward Bates, Project Manager 
U.S. Environmental Protection Agency 
National Risk Management Research Laboratory 
26 West Martin Luther King Drive 
Cincinnati, Ohio 45268 
Telephone: (513)569-7774 
Fax: (513)569-7676 


11 



Section 2 

Technology Applications Analysis 


This section of the ITER describes the general applicability 
of the CWS technology to contaminated waste sites. The 
analysis is based primarily on the SITE CWS treatability 
study and demonstration results. A detailed discussion 
of the treatability study and demonstration results is 
presented in Section 3.0 of this report. An article containing 
a constructed wetlands case study is presented in 
Appendix B. 

2.1 Applicable Wastes 

Constructed wetlands have been demonstrated to be 
effective in removing organic, metal, and nutrient elements 
including nitrogen and phosphorus from municipal 
wastewaters, mine drainage, industrial effluents, and 
agricultural run-off. The technology is waste-stream 
specific, requiring characterization of all organic and 
inorganic constituents. 

Because constructed wetlands can treat a wide variety of 
wastes, they vary considerably in their design. Constructed 
wetlands can be simple, single-cell systems, such as the 
two cells evaluated during this demonstration, or complex 
multicell or multicomponent systems. Complex constructed 
wetlands may include multiple wetland cells in series, 
anoxic limestone drains, marshes, ponds, and rock filters. 
Constructed wetlands tested in the eastern U.S. to 
remediate slightly acidic coal mine drainage have 
incorporated an anoxic limestone drain to provide alkalinity, 
followed by a holding pond, a constructed wetland, a 
shallow marsh, and finally a rock filter. The holding pond 
and wetland promote precipitation of iron hydroxides, 
while the marsh and rock filter remove manganese and 
suspended solids. Constructed wetlands design criteria 
are discussed in detail in an article by Gusek and Wildeman 
(1995). 

The results of the CWS demonstration (see Section 3.0) 
suggest the primary metals removal mechanisms are not 
identical within the upflow and downflow wetland cells. 


In the upflow cell, biological sulfate reduction appeared 
to be the primary zinc removal mechanism. Metals 
shown to be removed by this process include cadmium, 
copper, iron, lead, nickel, silver, and zinc (PRC 1995). In 
addition, biological sulfate reduction may also remove 
cobalt, mercury, and molybdenum contamination. In the 
downflow cell, chemical precipitation appeared to be the 
primary zinc removal mechanism. Because of the 
numerous geochemical species and complex equilibria 
involved in wetlands treatment of mine drainage, it is often 
difficult to predict which metals will precipitate. 
An equilibrium aqueous geochemical wetlands model 
(MINTEQ.AK) has been developed to help predict metal 
removal by constructed wetlands (Klusman 1993). 

2.2 Factors Affecting Performance 

Because CWS designs are so diverse, the number of 
parameters affecting their operation is also large. In the 
discussion that follows, the performance factors described 
pertain to this demonstration CWS (anaerobic compost) 
or to similar systems treating metal-contaminated mine 
drainage. These performance factors may or may not be 
relevant to constructed wetlands designed to treat organic 
or inorganic (nonmetal) contamination. Several factors 
influenced the performance of the two demonstration 
CWS. These factors can be grouped into three categories: 
(1) mine drainage characteristics, (2) operating parameters, 
and (3) compost degradation. 

2.2 .1 Mine Drainage Characteristics 

The CWS technology is capable of treating a range of 
contaminated waters containing heavy metals. However, 
the effectiveness of a CWS can be reduced as 
contaminants in high concentrations precipitate and clog 
the system prematurely. Often, contaminated coal mine 
drainages in the eastern U.S. contain elevated 
concentrations of iron or aluminum. When the pH of these 
drainages is raised during wetland treatment, iron and 


12 



aluminum hydroxides can form and precipitate (Hedin and 
others 1994). 

These precipitates can lead to a loss of permeability or a 
gradual filling of the wetland. Because sulfate-reducing 
bacteria cannot survive in low pH environments, low pH 
mine drainage can also affect the ability of the biological 
sulfate reduction wetland to remove contaminants. The 
oxidation/reduction potential (ORP) of the mine drainage 
may also affect the performance of the constructed 
wetland. However, the extent of the ORP effect is 
unknown. 

2.2.2 Operating Parameters 

The operating parameters that can be adjusted during the 
treatment process include the flow rate and hydraulic 
residence time of water within the wetland. In general, 
the selection and design for the hydraulic residence time 
is a function of the rate of metal loading. A hydraulic 
residence time of 50 to 100 hours was found to work well 
in the biological sulfate reduction reactors used during the 
short-term CWS treatability study (Figure 3). 

The residence time in the upflow and downflow cells 
during the demonstration was calculated at between 50 
and 60 hours. The calculation was based on the substrate 
volume of the wetlands, the percent moisture of the 
substrate (generally, 50 to 65 percent with 50 percent used 
in the calculation), and a flow rate of 7 gpm. 

Maintaining proper hydraulic residence times is one of the 
most important factors in successful wetlands treatment. 
In biological-based systems, a short residence time may 
not allow metals to precipitate and be filtered out by the 
wetland or may expose the bacteria to inhibitory levels of 
metal contaminants. Both may result in lower metal 
removal rates. In chemical precipitation systems, 
compounds that precipitate slowly may not be removed to 
the same extent as rapidly precipitating compounds. 

Chemical amendments, such as alkalinity or nutrients, are 
also examples of parameters that can be adjusted during 
the wetland treatment process. Alkalinity may be added 
via an anoxic limestone drain or directly to the mine 
drainage as lime. Nutrients could also be added directly 
to the mine drainage or applied to the ponded surface 
water of downflow cells. Neither alkalinity nor nutrients 
was added to the SITE demonstration CWS. 


2.2.3 Compost Performance 

Compost performance depends on the compost materials 
used and the characteristics of the mine drainage. When 
using manure compost, the metals concentrations of the 
drainage, the nutrient concentrations in the compost, and 
gradual breakdown and compaction of the compost 
materials are the most important factors controlling compost 
effectiveness. Of these factors, substrate breakdown 
and compaction that leads to a loss ofhydraulic conductivity 
is probably the most important factor. The breakdown of 
the complex biological polymers to smaller compounds by 
fermentative bacteria gradually destroys the structural 
intensity of the compost and leads to compaction. One 
way to extend substrate lifetime is to include materials that 
are degraded at a moderate rate. Based on the loss of 
nutrients and hydraulic conductivity in the upflow CWS, 
the wetland compost material is expected to last 4 to 
5 years before becoming ineffective. 

The accumulation of metals within the constructed wetlands 
may eventually cause the compost material to become a 
hazardous waste, substantially decreasing the number of 
compost disposal options and increasing treatment costs. 

However, after 4 years of near-continuous operation of 
the demonstration CWS, neither cell’s compost material 
developed hazardous characteristics based on thresholds 
defined in 40 Code of Federal Regulations (CFR) 
Part 261.24. However, the primary contaminant in the 
Burleigh Mine Drainage, zinc is not a TCLP analysis 
parameter. 

2.3 Site Characteristics 

Site characteristics are important when considering CWS 
technology because they can affect system application. 
All characteristics should be considered before selecting 
the technology to remediate a specific site. Site-specific 
factors include support systems, site area and preparation, 
site access, climate, hydrology, utilities, and the availability 
of services and supplies. 

2. 3.1 Support Sys terns 

If on-site facilities are not already available, a small 
storage building equipped with electricity may be desirable 
near the treatment system. The on-site building could be 
used for storing operating and sampling equipment (tools, 
field instrumentation, and health- and safety-related gear) 
and providing shelter for sampling personnel during 
inclement weather. The building may also be used for 
calibrating field equipment for system monitoring. 


13 





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Figure 3. Flow rates measured for effluent cells. 







2.3.2 Site Area, Preparation, and Access 

Constructed wetlands typically require a larger level 
area compared to other treatment options. The results 
of this investigation suggest that a 50-60 hour hydraulic 
residence time is near the lower limit required of these 
systems to provide consistent metal removal. Researchers 
in this field have suggested that longer residence times 
ranging from 75 to 150 hours may be required for long¬ 
term metal removal (Dr. Ronald Klusman and Dr. Richard 
Gammons, personal communications) The depth of the 
compost in the demonstration CWS cells was 4 feet. The 
maximum depth of compost that can be used while 
maintaining treatment effectiveness is unknown. 
Consequently, some sites may require extensive grading 
and leveling to allow construction of a CWS. Depending 
on the site, grading and leveling may be cost prohibitive. 

Piping or other mechanisms for conveying mine drainage 
to the wetlands is also necessary. In addition, a relatively 
constant rate of flow is desired to keep the wetlands 
active. Thus, site conditions may require a mine 
drainage collection, storage, and distribution structure. 

Furthermore, an upflow constructed wetland may require 
that the mine drainage distribution network include a dam 
or pump to maintain sufficient hydraulic head to force 
the mine drainage through the compost. Also, piping is 
required to bypass flow around the wetland. This bypass 
piping or conveyance should be oversized to manage 
200 to 300 percent of the predicted maximum mine 
drainage discharge. 

Access roads for heavy equipment (excavation and 
hauling) are required to install, operate, and maintain a 
CWS. 

2.3.3 Climate 

The climate at potential constructed wetland sites can be 
a limiting factor. Extended periods of severe cold, 
extreme hot and arid conditions, and frequent severe 
storms or flooding will affect system performance. 
Extreme cold can freeze portions of the wetland resulting 
in channeling of the mine drainage through the substrate, 
thus, reducing the hydraulic residence time. In addition, 
cold temperatures may reduce microbial activity or 
populations. Reductions in hydraulic residence time and 
microbial activity will both lessen the ability of the 
constructed wetland to remove metals and may require it 
to be oversized. The large water surface areas and plant 
life associated with wetlands enhance evaporation and 
evapotransportation. A constructed wetland in a hot and 


arid climate may periodically dry up at a site with low 
water flow rates. If the wetland design does not consider 
cyclical periods of wet and dry, it may be less effective 
during the wet periods. Constructing wetlands in areas 
with frequent flooding or severe storms can lead to 
hydraulic overloading or washout of substrate materials. 
The engineering controls required to overcome these 
climatic or geographic limitations may eliminate the low 
cost and low maintenance advantages that make 
constructed wetlands appealing. 

2.3.4 Utilities 

The CWS is a passive treatment technology, so utilities 
are not required to operate the system. However, in some 
situations electricity for pumps or on-site analytical 
instrumentation may be desirable. In remote areas, an on¬ 
site storage building should be provided if possible. A 
telephone connection or cellular phone is required for 
operating and sampling personnel to contact emergency 
services if needed and for routine communications. 

2.3.5 Services and Supplies 

The main services required by the CWS are periodic 
adjustment of system flow rates, cleanout of effluent 
piping, and the removal and replacement of compost 
materials. During the CWS demonstration, flow rate 
adjustments were required every 3 to 6 months, and effluent 
piping cleanout was conducted once. However, both 
CWS demonstration cells were operated from a single v- 
notch weir and the flow diverted to the cells. The 
frequency of flow adjustment would be lower if each cell 
had been constructed with its own weir. The time 
between changeout of wetland compost depends on the 
chemical constituents of the influent water, the 
configuration and capacity of the constructed wetland, 
and the preferred method of disposal. The compost 
lifetime, estimated from nutrient loss and the development 
of short circuiting during this demonstration is estimated to 
be 4 to 5 years. 

2.4 Availability, Adaptability, and 
Transportability of Equipment 

The components of a simple CWS are generally available 
locally. The components include standard construction 
materials for the structure of the wetland cells, liner 
materials available from several sources, and compost 
materials, the type of which will depend on the contaminants 
in the mine drainage. The most suitable compost for a 
given application can be identified during a treatability 
study using materials available locally. 


15 




2.5 Material Handling Requirements 

The CWS generates spent compost material. Substrate 
material will require testing to evaluate disposal options. 
Depending on the disposal option, dewatering or other 
pretreatment may be necessary prior to shipment for off¬ 
site disposal. Depending on regulatory requirements, the 
effluent water generated during dewatering may also 
require additional treatment prior to discharge. 

Some CWS compost materials may contain high levels of 
water-soluble nitrogen or phosphorus compounds. These 
compounds can be readily leached from the fresh compost 
during startup of the constructed wetland. Thus, the CWS 
effluent at startup may require treatment to reduce or 
remove excess nitrogen or phosphorous. Treatment may 
include land application, if permitted, or effluent collection 
for subsequent recycling through the CWS. 

2.6 Personnel Requirements 

Wetlands construction and compost replacement require 
heavy equipment operators, laborers, and a construction 
supervisor. After the CWS is installed, personnel 
requirements include a sampling team and personnel to 
adjust system flow rates. Sampling personnel should be 
able to collect water and substrate samples for laboratory 
analysis and measure field parameters using standard 
instrumentation. 

All personnel should have completed an Occupational 
Safety and Health Administration (OSHA) initial 40-hour 
health and safety training course with annual 8-hour 
refresher courses, if applicable, before constructing, 
sampling, replacing compost, or removing a constructed 
wetland at hazardous waste sites. They should also 
participate in a medical monitoring program as specified 
under OSHA requirements. 

2.7 Potential Community Exposures 

Fencing and signs should be installed around a CWS to 
restrict access to the system for both humans and wildlife. 
The potential routes of exposure include the mine drainage 
or waste stream, the compost material, and the CWS 
effluent. The actual exposure risk depends on the 
constituents of the specific waste being treated and the 
effectiveness of the treatment. 

The CWS may also generate low concentrations of 
hydrogen sulfide gas, depending on the time of year and 
the biological activity of the CWS. Odors caused by 


hydrogen sulfide and volatile fatty acids from the decaying 
manure may be a nuisance to a local community. 

2.8 Evaluation of Technology Against 
RI/FS Criteria 

EPA has developed nine evaluation criteria to fulfill 
the requirements of the Comprehensive Environmental 
Response, Compensation, and Liability Act (CERCLA), 
as well as additional technical and policy considerations 
that have proven important for selecting potential remedial 
alternatives. These criteria serve as the basis for 
conducting bench-scale testing during the remedial 
investigation (RI) at a hazardous waste site, for conducting 
the detailed analysis during the feasibility study (FS), and 
for subsequently selecting an appropriate remedial action. 
Each SITE technology is evaluated against the nine EPA 
criteria because these technologies may be considered as 
potential remedial alternatives. The nine evaluation criteria 
are: 

• Overall protection of human health and the 
environment 

• Compliance with applicable or relevant and 
appropriate requirements (ARAR) 

• Long-term effectiveness and permanence 

• Reduction of toxicity, mobility, or volume 

• Short-term effectiveness 

• Implementability 

• Cost 

• State acceptance 

• Community acceptance 

Table 1 presents the results of this evaluation for the 
CWS. The demonstration results indicate the up flow 
CWS can provide short-term protection of the environment; 
reduces contaminant mobility, toxicity, and volume; is cost 
effective; implementable, and is an acceptable remedy to 
the community and state regulators. However, neither 
CWS cell tested in this demonstration, provided long-term 
effectiveness. This in part is the result of low zinc 
discharge requirements (200 pg/L) at the demonstration 
site. Other sites may have less strict discharge 
requirements. In addition, the upset condition resulting 
from the high flow event also contributed to the lack of 
long-effectiveness particularly in regards to the upflow 
cell. 


16 



Table 1. Evaluation of CWS Treatment Versus RI/FS Criteria 


Criterion 


Discussbn 


1. Overall Protection of Human Health 
and the Environment 


2. Compliance with Applicable or 
Relevant and Appropriate 
Requirements (ARAR) 


3. Long-Term Effectiveness and 
Permanence 


4. Short-term Effectiveness 


5. Reduction of Toxicity, Mobility, or 
Volume of contaminates through 
Treatment 

6. Implementability 


7. Cost 


8. Community Acceptance 


9. State Acceptance 


As tested, the CWS provided only short-term 
effectiveness. Indifferentcircumstances.theCWS may 
provide short- and tong-term protection by removing 
mine drainage contaminants. 

Substrate is a recycled product, not mined or 
manufactured. 

Wetland effluentdischarge may require compliance with 
Clean Water Act regulations. 

Substrate disposal may require compliance with RCRA 
regulations. 

CWS treatment removes contamination from mine 
drainage, but may not meet low-level discharge 
requirements. 

Use of CWS treatment with other technologies may be 
effective in meeting low-level discharge requirements. 

Presents few short-term risks to workers, community, or 
wildlife. 

Minimal personal protective equipment required for 
operators. 

CWS treatment reduces contaminant mobility, toxicity, 
and volume. 


Generally a passive treatment system, but can be 
active. 

Construction uses standard material and practices 
common in the industry. 

Construction cost of full-scale (50gpm) system is 
estimated at approximately $290,000. 

O&M of full-scale CWS system is estimated to be 
$57,000 per year. 

The public usually views the technology as a natural 
approach to treatment; therefore, the public generally 
accepts this technology. 

CDPHE found the technology shows promise for 
treating AMD; however, based on constraints at the 
Burleigh site, including the cold climate and proximity to 
town, CDPHE recommended not implementing a full- 
scale, permanent system at the site. 

Colorado Division of Minerals has built several CWSs to 
treat AMD. 


17 






2.9 Potential Regulatory Requirements 


This section discusses specific environmental regulations 
pertinent to operation of a CWS, including the transport, 
treatment, storage, and disposal of wastes and treatment 
residuals, and analyzes these regulations in view of the 
demonstration results. State and local regulatory 
requirements, which may be more stringent, must also be 
addressed by remedial managers. 

ARARs include the following: (1) CERCLA; (2) the 
Resource Conservation and Recovery Act (RCRA); 
(3) the Clean Water Act; and (4) OSHA regulations. 
These four general ARARs are discussed below; specific 
ARARs must be identified by remedial managers for each 
site. 

2.9.1 Comprehensive Environmental 
Response , Compensation , and 
Liability Act 

CERCLA, as amended by SARA, authorizes the federal 
government to respond to releases or potential releases of 
any hazardous substance into the environment, as well as 
to releases of pollutants or contaminants that may present 
an imminent or significant danger to public health and 
welfare or the environment. 

As part of the requirements of CERCLA, EPA has 
prepared the National Oil and Hazardous Substances 
Pollution Contingency Plan (NCP) for hazardous substance 
response. The NCP, codified at 40 CFR Part 300, 
delineates methods and criteria used to determine the 
appropriate extent of removal and cleanup for hazardous 
waste contamination. 

SARA amended CERCLA and directed EPA to: 

• Use remedial alternatives that permanently and 
significantly reduce the volume, toxicity, or mobility 
of hazardous substances, pollutants, or contaminants. 

• Select remedial actions that protect human health 
and the environment, are cost-effective, and involve 
permanent solutions and alternative treatment or 
resource recovery technologies to the maximum 
extent possible. 

• Avoid off-site transport and disposal of untreated 
hazardous substances or contaminated materials when 
practicable treatment technologies exist (Section 
121[b]). 

In general, two types of responses are possible under 
CERCLA: removals and remedial actions. The CWS 


technology is likely to be part of a CERCLA remedial 
action. Remedial actions are governed by CERCLA as 
amended by SARA. As stated above, these amendments 
promote remedies that permanently reduce the volume, 
toxicity, and mobility of hazardous substances, pollutants, 
or contaminants. 

On-site remedial actions must comply with federal and 
state ARARs. ARARs are identified on a site-by-site 
basis and may be waived under six conditions: (1) the 
action is an interim measure, and the ARAR will be met 
at completion; (2) compliance with the ARAR would pose 
a greater risk to human health and the environment than 
noncompliance; (3) it is technically impracticable to meet 
the ARAR; (4) the standard of performance of an ARAR 
can be met by an equivalent method; (5) a state ARAR 
has not been consistently applied elsewhere; and (6) 
ARAR compliance would not provide a balance between 
the protection achieved at a particular site and demands 
on the Superfund for other sites. These waiver options 
apply only to Sup^rfund actions taken on site, and 
justification for the waiver must be clearly demonstrated. 

2.9.2 Resource Conservation and 
Recovery Act 

RCRA, an amendment to the Solid Waste Disposal Act 
(SWDA), was enacted in 1976 to address the problem of 
safe disposal of the enormous volume of municipal and 
industrial solid waste generated annually. RCRA 
specifically addressed the identification and management 
of hazardous wastes. The Hazardous and Solid Waste 
Amendments of 1984 (HSWA) greatly expanded the 
scope and requirements of RCRA. 

The presence of RCRA-defined hazardous waste 
determines whether RCRA regulations apply to the 
CWS technology. RCRA regulations define and regulate 
hazardous waste transport, treatment, storage, and disposal. 
Wastes defined as hazardous under RCRA include 
characteristic and listed wastes. Criteria for identifying 
characteristic hazardous wastes are included in 40 CFR 
Part 261 Subpart C. Listed wastes from nonspecific and 
specific industrial sources, off-specification products, spill 
cleanups, and other industrial sources are itemized in 40 
CFR Part 261, Subpart D. 

The CWS demonstration treated mine discharge water 
from the Burleigh Tunnel, which is included in the Clear 
Creek/Central City Superfund site. The manure compost 
was tested regularly to determine whether it would become 
a hazardous waste during the demonstration. The concern 


18 



was that sorption and precipitation of metals could cause 
the substrate to become a hazardous waste, thus restricting 
options and increasing cost for material disposal. The 
substrate did not exhibit the characteristics of hazardous 
waste after nearly 4 years of operation. 

2.9.3 Clean Water Act 

The objective of the Clean Water Act is to restore and 
maintain the chemical, physical, and biological integrity of 
the nation’s waters. To achieve this objective, effluent 
limitations of toxic pollutants from point sources were 
established. Wastewater discharges are most commonly 
controlled through effluent standards and discharge permits 
administered through the National Pollutant Discharge 
Elimination System (NPDES) by individual states with 
input from the federal EPA. Under this system, discharge 
permits are issued with limits on the quantity and quality 
of effluents. These limits are based on a case-by-case 
evaluation of potential environmental impacts and on 
wasteload allocation studies aimed at distributing discharge 
allowances fairly. Discharge permits are designed as an 
enforcement tool with the ultimate goal of achieving 
ambient water quality standards (Metcalf and Eddy 1979). 

NPDES permit requirements must be evaluated for each 
CWS when the effluent water is discharged into a 
waterway or water body. The requirements and standards 
that must be met in the effluent for each CWS will be 
based on the waterway or water body into which the CWS 
discharges. The effluent limits will be established through 
the NPDES permitting process by the state in which the 
CWS is constructed and by EPA. 

CDPHE has identified stream standards for Clear Creek 
at the Burleigh Tunnel discharge. Table 2 provides these 
standards for both low- and high-flow conditions. The 
zinc standard for both low- and high-flow conditions is 200 
pg/L in the receiving stream (Clear Creek). In order to 
met this standard, the discharge from Burleigh Tunnel 
must contain less than 13,650 pg/L zinc under low-flow 
conditions and 65,700 pg/L under high-flow conditions. 

2.9.4 Occupational Safety and Health Act 

CERCLA remedial actions and RCRA corrective actions 
must be conducted in accordance with OSH A requirements 
detailed in 29 CFR Parts 1900 through 1926, especially 
Part 1910.120, which provides for health and safety of 
workers at hazardous waste sites. On-site construction at 
Superfund or RCRA corrective action sites must be 
conducted in accordance with 29 CFR Part 1926, which 


provides safety and health regulations for construction 
sites. State OSHA requirements, which may be 
significantly stricter than federal standards, must also be 
met. 

Construction and maintenance personnel and sampling 
teams for the Burleigh Tunnel CWS demonstration all 
met the OSHA requirements for hazardous waste sites. 
For most sites, the minimum personal protective equipment 
(PPE) required would include gloves, hard hats (during 
construction), steel toed boots, and eye protection. 
Additional PPE may be required during summer or winter 
months to protect against extreme temperatures. 

2.10 Limitations of the Technology 

Land required for constructed wetland systems is typically 
extensive compared to conventional treatment systems. 
Thus, in areas with high land values, a constructed 
wetland treatment system may not be appropriate. Land 
availability relatively close to the source of contaminated 
water is preferred to avoid extended transport. 

The climate at potential constructed wetland sites can also 
be a limiting factor. Extended periods of severe cold, 
extreme heat, arid conditions, and frequent severe storms 
or flooding can result in performance problems. 
Contaminant levels in treated and discharged water can 
vary in response to variations of influent volumes and 
chemistry. This may also be a limiting factor if there is no 
tolerance in contaminant level discharge requirements. 


19 






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21 





Section 3 

Treatment Effectivemess 


The following sections discuss the treatment effectiveness 
of the CWS demonstration in Silver Plume, Colorado. 
The discussion includes abackground section, a review of 
the demonstration, demonstration methodology, site 
demonstration results, and demonstration conclusions. 

3.1 Background 

The Burleigh Tunnel is located approximately 50 miles 
west of Denver in the Georgetown-Silver Plume mining 
district (Figure 1). The Georgetown-Silver Plume mining 
district occupies an area of about 25 square miles 
surrounding the towns of Silver Plume and Georgetown. 
In general, the period of significant silver production in the 
area commenced in 1872, reached a peak in 1894, and 
gradually declined after. Mining in the district increased 
briefly during World Wars I and II, when many old mines 
were reopened and considerable amounts of lead and zinc 
were mined from old stopes, dumps, and wastes left from 
the silver mining boom. 

The Burleigh Tunnel drains a group of mines on Sherman 
and Republican mountains. Many of these mines intercept 
shallow groundwater migrating through fractures in the 
rock or surface water collected by stopes. The intercepted 
waters are transported through the mines and are eventually 
discharged through the Burleigh Tunnel. The Burleigh 
Tunnel discharge contains elevated levels of zinc, typically 
between 45 and 65 mg/L. However, greater than normal 
precipitation during the spring of 1995 mobilized a large 
amount of zinc and increased zinc concentrations within 
the drainage to 109 mg/L. Burleigh Tunnel discharge 
rates are generally between 40 to 60 gpm and increase to 
100 to 140 gpm during spring runoff. The elevated levels 
of zinc and significant flow rates combine to make the 
Burleigh Tunnel a major source of zinc to Clear Creek. 
Because of the large amount of zinc being discharged to 
Clear Creek and the potential impact of the zinc on the 
Clear Creek fishery, the drainage from the Burleigh 


Tunnel was included in the Clear Creek/Central City 
Superfund site. 

The elevation of the Burleigh Tunnel is 9,152 feet, and the 
climate is typical of mountainous alpine regions in Colorado. 
Summers are short and cool and winters are long and cold. 
Strong eastward, down-valley winds are typical during the 
winter months. Winds are lighter during the summer 
months and occasionally blow westward, up the valley. 
Snow accumulation during the winter months in the 
immediate area of the tunnel is usually not significant due 
to the open, south-facing exposure of the hillside and high 
winds. Snow accumulation at higher elevations in more 
sheltered areas is significant, with some snow fields 
persisting until late summer. The average annual 
temperature is approximately 43.5 degrees Fahrenheit 
(°F), with a mean minimum of 31 °F and a mean maximum 
of 55.9°F. The average annual precipitation is 15.14 
inches. 

3.2 Review of SITE Demonstration 

The SITE demonstration was divided into three phases: 
(1) CWS treatability study; (2) CWS technology 
demonstration; and (3) site demobilization. These activities 
are reviewed in the following sections, which also discuss 
variations from the work plan and the CWS performance 
during the technology demonstration phase. 

3.2 .1 Treatability Study 

A treatability study was conducted at the Burleigh Tunnel 
between June 18,1993, and August 12,1993. The goal of 
the treatability study was to show that bacterial sulfate 
reduction could remove zinc from the low-sulfate mine 
drainage from the Burleigh Tunnel and to estimate levels 
of zinc reduction that could be expected by CWS treatment. 
The treatability study involved the construction, operation, 
and sampling of two bioreactors. Each bioreactor was 


22 



filled with a mixture of composted manure (96 percent) 
and alfalfa hay (4 percent), the same substrate that was 
to be used in the CWS demonstration treatment cells. 
Both reactors used an upflow configuration, in which 
Burleigh Tunnel drainage entered the bioreactors from 
the bottom and was forced to flow up through the substrate. 
The small bioreactor was 4 feet tall and 22 inches in 
diameter and held approximately 60 gallons of compost 
and water. The large bioreactor was 8 feet tall and 22 
inches in diameter and held approximately 130 gallons of 
compost and water. The lower 6 inches of each bioreactor 
was filled with gravel to support inlet piping and minimize 
channeling. Peristaltic pumps were used to establish a 
flow rate of 20 to 30 milliliters per minute for the small 
bioreactor and 50 to 60 milliliters per minute for the large 
bioreactor. The flow rates for the bioreactors were set to 
provide an estimated hydraulic residence time of 50 to 
100 hours. 

The results of the treatability study indicated that after 
8 weeks of operation, both bioreactors achieved removal 
efficiencies of 99 percent for zinc and similar efficiencies 
for cadmium and manganese. Zinc was the major metal 
of concern for the Burleigh Tunnel drainage. Sorption of 
metals in the substrate is believed to be the dominant 
removal process during the first 1 to 2 weeks of bioreactor 
operation. After this brief period of sorption, biological 
sulfate reduction apparently became the primary metal 
removal process in the bioreactors. Results of sulfate- 
reducing bacteria counts and sulfate and sulfide analyses 
indicated that a large population of sulfate-reducing 
microorganisms was active in the system. The results 
supported the theory that the bacteria reduce sulfate in the 
water to hydrogen sulfide ions, which react with dissolved 
metals to produce insoluble metal sulfides. The results 
indicated that the Burleigh Tunnel drainage contains a 
sufficient concentration of sulfate to promote metal removal 
by microbial sulfate reduction. Compost sample results 
from both bioreactors indicated that the compost 
accumulated metals and sulfide but did not become a 
reactive or hazardous waste after 8 weeks of operation. 

3.2.2 Technology Demonstration 

Site preparation requirements for the CWS demonstration 
were minimal because of previous mining and treatability 
study activities. Moreover, the area surrounding the 
Burleigh Tunnel adit is level and required only minor 
grading to install the two CWS treatment cells. Construction 
of the CWS treatment cells and all drainage conveyances 
was the responsibility of the developer (CDPHE). 


The demonstration evaluated two treatment cells that 
differed only in flow configuration, one upward and the 
other downward. The demonstration evaluated the ability 
of each cell to remove zinc and other metals from the 
Burleigh Tunnel mine drainage without pretreatment. 
Efforts were made to maintain constant flow rates; 
however, flow rates did vary. In addition, several events 
resulted in brief interruptions of flow to the cells. 
Approximately 12.7 million gallons of water from the 
Burleigh Tunnel were passively treated by the upflow 
constructed wetland cell and 11 million gallons by the 
downflow CWS over the 46-month demonstration. 
Figure 3 shows the flow rates measured for both wetland 
cell effluents during the demonstration. 

Throughout the demonstration, mine drainage influent and 
wetlands system effluent samples were collected for 
analysis of total metals, anions, total suspended solids 
(TSS), and total organic carbon (TOC). In addition, 
wetlands substrate samples were collected monthly for 
sulfate-reducing bacteria analysis and quarterly for analysis 
of total metals, acid-volatile sulfides (AV S), and toxicity 
characteristic leaching procedure (TCLP) metals. The 
substrate samples were analyzed to evaluate the 
effectiveness of the treatment system in sequestering 
zinc, to assess the tendency of the substrate to become a 
hazardous waste, and to estimate the role of sulfate- 
reducing bacteria within the wetlands substrate. 

3.2.3 Operational and Sampling Problems 
and Variations from the Work Plan 

The CWS experienced several operational problems during 
the demonstration. Some of these problems resulted in 
changes to the schedule and sampling events. Problems 
encountered and resolutions effected during the 
demonstration are described below. 

• The upflow cell froze in December 1993 and remained 
frozen until the middle of February 1994. The cell 
froze because flow to the cells was interrupted when 
the dike within the Burleigh Tunnel collapsed. The 
dike was quickly repaired; however, as a result of 
the cold conditions and the lack of flow to the cells, 
the upflow cell froze to a depth of 18 inches. A 
livestock water heater and a steam cleaner were 
used to thaw the cell so that flow through the cell 
could be maintained. The freezing of the upflow cell 
delayed the start of the demonstration by 1 month. 
In order to prevent the upflow cell from freezing 
during the winter of 1995, straw bales were placed 
on top of the cell to provide insulation from the cold. 

• The insulation provided by the straw bales maintained 
the wetland water temperatures consistent with 


23 



influent values and the up flow cell effluent piping did 
not freeze. 

• The 1995 spring runoff was exceptionally high, and 
more flow was channeled to the CWS than the 
wetlands were designed to handle. More than 
20 gpm were flowing through the upflow cell for a 2- 
week period in early June 1995. CDPHE responded 
to the flooding by installing a 6-inch bypass pipe to 
carry overflow from the influent weir around the 
wetlands Once installed, the bypass allowed flow 
rates to be returned to 7 gpm for each cell. However, 
CDPHE had not removed the straw bales insulating 
the upflow cell before the spring runoff began, and 
the straw bales became saturated. The weight of 
the saturated straw compressed the substrate, 
reducing the flow within the upflow cell to less than 
1 gpm. The straw bales were removed from the 
upflow cell, and flow was restored to the cell within 
a week. 

• In late November 1994, a large block of rock, 
roughly 10 feet by 10 feet, fell from the hillside and 
rolled onto a comer of the upflow CWS cell. The 
rock appeared to have depressed the effluent 
accumulation network and created a high spot in the 
piping at the collection point to the effluent weir. 
The high point in the piping may have resulted in the 
collection of precipitated metal sulfides in the piping, 
causing a flow restriction. 

• During the summer and fall of 1994 and 1995, the 
effluent flowrate from the downflow cell could not 
be maintained at 7 gpm. It was not clear if biological 
surface growth, chemical precipitation in the cell, or 
settling and compaction of fine particles in the 
substrate was responsible for the decreased cell 
permeability. 

• Several substrate sampling techniques were proposed 
for the demonstration, including polyethylene dipper 
and sediment core samplers. Both techniques 
appeared to be equally effective; however, the dippers 
were determined to be preferable. The dippers 
were selected because they were inexpensive and 
could be dedicated to each sampling cell, reducing 
the number of equipment blank samples required 
during the demonstration. 

3.2.4 Site Demobilization 

The demonstration-scale wetland was removed by 
CDPHE at the end of the demonstration. Wetland removal 
entailed: 

• Removal and disposal of the wetland substrate 

Filling the wetland cells with site materials 
Filling or removal of wetland weirs 


• The CWS demonstration substrate was not a 
hazardous material, and potential disposal options 
included: 

Disposal at a municipal landfill 
Disposal in landfill biobeds (compost piles) 
Mixing with site mining waste rock and soil to 
provide needed organic matter 
Reuse in an interim ponded wetland 


• The CWS Demonstration substrate was disposed of 
in a nearby municipal landfill 

3.3 Demonstration Methodology 

The primary objectives of the CWS technology 
demonstration were to (1) measure the reduction of zinc 
in Burleigh Tunnel drainage resulting from the CWS 
treatment with respect to cell configuration and seasonal 
variation (temperature); (2) assess the toxicity of the 
Burleigh Tunnel drainage; (3) characterize the toxicity 
reduction resulting from treatment of the drainage by the 
CWS; and (4) estimate toxicity reductions in the stream 
(Clear Creek) receiving the Burleigh Tunnel drainage. In 
addition, secondary objectives of the demonstration 
included: 

• Estimating the metal removal capacity (lifetime) of 
the substrate, including the effect of treatment cell 
flow configuration. The results of influent and 
effluent metal analyses, CWS flow rate data, and 
TCLP metal analysis were compared to substrate 
metal accumulation estimates to evaluate the removal 
capacities of each CWS treatment cell, "lhe TCLP 
metals analysis was used because the substrate 
could become a hazardous waste before its metal 
removal capabilities were exhausted. Replacing the 
substrate before it becomes a hazardous waste was 
determined to be the most cost-effective solution. 

• Estimating the extent to which sulfate-reduction 
processes within the CWS are responsible for the 
removal of zinc from the drainage. Substrate was 
analyzed for sulfate-reducing bacteria and acid- 
volatile sulfides to estimate the extent to which sulfate- 
reduction processes are removing zinc from the 
drainage. The approximate number of sulfate- 
reducing bacteria was correlated to metal removal 
efficiencies as part of the determination. In addition, 
the accumulation of AVS in the substrate was 
compared to metal loading in the treatment cells to 
determine trends. Furthermore, the AVS analyses 
included an analysis of zinc to verify that the metal 
sulfides accumulating in the CWS were zinc sulfides. 
Previous investigations suggested that AVS analyses 
were indicative of metal sulfide accumulation 
attributed to sulfate-reducing bacteria (Reynolds 
1991). 


24 



• Evaluating the impact of the CWS effluent on Clear 
Creek. Clear Creek samples were analyzed for total 
metals, TSS, total dissolved solids (TDS), TOC, 
nitrate, and phosphate. Results of the stream analyses 
were compared to CWS effluent analyses to assess 
the effect of CWS effluent on Clear Creek. Clear 
Creek samples were collected upstream and 
downstream of the CWS outfall. 

• Estimating the capital and operating costs of the 
CWS. 

Critical parameters are the data required to meet the 
primary objectives. The primary critical parameters were 
influent and effluent analyses for zinc (total), and toxicity 
testing with fathead minnows (Pimephalus promelas) and 
water fleas (Ceriodaphnia dubia). 

Noncritical parameters are data required to address 
secondary objectives of the demonstration. Secondary 
objectives provide useful information to potential technology 
users but are not critical to evaluate the technology. The 
noncritical parameters of the C W S demonstration included: 

• Total metals, nitrate and phosphate analysis of the 
Burleigh Tunnel drainage and CWS effluents 

• Metal loading, metal accumulation, and TCLP metals 
in CWS substrate samples 

• Sulfate-reducing bacteria counts and AVS 
accumulation in CWS substrate samples 

• Clear Creek samples for total metals, TDS, TSS, 
TOC, biochemical oxygen demand (BOD), and 
aquatic toxicity 

• Construction, operation, maintenance, substrate 
disposal, and miscellaneous costs 

3.3.1 Testing Approach 

In general, the testing approach of the demonstration 
incorporated the collection and analysis of wetland influent 
and effluent samples every 2 weeks for a period of 
20 months. Monthly sampling was conducted for the 
remainder of the nearly 4-year demonstration. The 
effluent zinc results for each sampling event were 
compared to influent data and a removal efficiency 
calculated. An initial 2-week interval was selected 
because it provided for 3 to 7 pore volumes of water to be 
passed through the CWS, assuming a hydraulic residence 
time of between 50 and 100 hours. In addition, the 2-week 
interval was chosen because several factors, such as 
precipitation or evaporation, could cause variation in the 
measured concentration of zinc in wetland effluent samples. 
By increasing the number of influent and effluent water 


samples, performance trends display better continuity, the 
effects of weather are reduced, and calculated removal 
efficiencies are expected to more closely reflect true 
values. Also, sampling intervals shorter than 2 weeks 
were not economically feasible considering the length of 
the demonstration. The initial 20-month schedule was the 
maximum time allowable for the demonstration. This time 
frame is allowed because the CWS is a biological 
technology and performance depended, in part, on primary 
substances and nutrients within the substrate. By allowing 
the system to operate for an extended period, results were 
expected to show a relationship (positive or negative) 
between declining nutrient concentrations in the substrate 
and CWS performance. 

The frequency of demonstration toxicity testing was 
limited to every 3 to 4 months due to budget considerations. 
Essentially, the sample collection and testing schedule 
was designed to evaluate toxicity reduction during periods 
of widely different zinc removal (different seasons) and 
critical periods for the receiving stream. 

3.3.2 Sampling, Analysis, and 
Measurement Procedures 

Mine drainage samples were collected from the influent 
weir, and CWS effluent samples were collected from the 
effluent weirs. Clear Creek samples were collected 
above and below the CWS outfall. Influent and effluent 
samples were analyzed for total recoverable zinc and 
toxicity (critical analyses), other metals, anions, TDS, 
TSS, and TOC (effluent only). These samples were 
collected at the frequency discussed in the previous 
section. 

Two substrate sampling points were located in each cell. 
Initially, substrate samples were collected monthly for 
sulfate-reducing bacteria analysis and quarterly for total 
metals, AVS, and TCLP metals analyses for a period of 
20 months. Quarterly and semi-annual sampling was 
conducted for the remainder of the demonstration. 
Substrate samples were collected from two locations 
within each cell, at approximately 1 to 2 feet below the 
wetland surface. 

Mine drainage, wetlands effluent, and substrate were 
analyzed for critical and noncritical parameters using the 
methods listed in Table 3. 

Field analyses included measurement of pH and 
conductivity for all aqueous samples. Eh for wetlands 
effluent samples, and dissolved oxygen for mine drainage 


25 




Table 3. CWS Demonstration Summary of Standard Analytical Methods and Procedures 


Parameter 

Sample Type 

Method Number 

Method Title 

Source 

Metals 

Aqueous and 
Substrate 

6010A,6020, 7470 

ICP, 1 CP/MS, or AA 

SW-846 1 

Sulfate 

Aqueous 

300.0 

Ion chromatography 

MCAWW2 

Fluoride 

Aqueous 

9056 

Ion chromatography 

SW-846 

Nitrate/Nitrite 

Aqueous 

353.2 and 354.1 

Various 

MCAWW2 

Chloride 

Aqueous 

300.0 

Ion chromatography 

MCAWW2 

Total and 

Aqueous 

365.3 

Various 

MCAWW 

Orthophosphate 

pH 

Aqueous 

9040 

Electrometric 

MCAWW 

TSS 

Aqueous 

160.2 

Gravimetric 

MCAWW 

TDS 

Aqueous 

160.1 

Gravimetric 

MCAWW 

TOC 

Aqueous 

9060 

Various 

SW-846 

Ammonia 

Aqueous 

350.1 

Various 

MCAWW2 

Alkalinity 

Aqueous 

310.1 

Various 

MCAWW2 

Sulfide 

Aqueous 

376.2 

Various 

MCAWW2 

Aquatic Toxicity 

Aqueous 

EPA SOPs 3 


EPA 5 

Acid Volatile Sulfide 

Substrate 

EPA Method 

Acid volatile sulfide 

EPA 1991 

(A VS) 

Sulfate reducing bacteria 

Substrate 

None 

Anaerobic deep tube 

CSM 3 

count 

Toxicity leaching 

Substrate 

1311 

ICP, ICP-MS or AA 

SW-846 

procedure 

Reactive sulfide 

Substrate 

EPA 4 

Titration 

SW-846 

Orthophosphate 

Substrate 

365.3 

Various 

MCAWW 

Sulfate 

Substrate 

300.0 

Various 

MCAWW 

Physical parameters 

Substrate 

Various3 

Various3 

ASTM 

Residence time 

Aqueous 

ND 

ND 

ND 

pH 

Aqueous 

SOP 3 12 


Tetra Tech 6 

Temperature 

Aqueous 

SOP 3 11 


Tetra Tech 6 

Dissolved oxygen 

Aqueous 

SOP 3 62 


Tetra Tech6 

Conductivity 

Aqueous 

SOP 3 99 


Tetra Tech6 


Notes: 


1 Test Methods for Evaluating Solid Wastes, Volumes IA- 1C: Laboratory Manual, Physical/Chemical Methods; and 
Volume II Field Manual. Physical/Chemical Methods, SW-846. 3d Edition. Office of Solid Waste and Emergency 
Response. U.S. Environmental Protection Agency (EPA). 1986. 

2 Methods for Chemical Analysis of Water and Wastes (MCAWW). EPA 600/4-79-020. Environmental Monitoring and 
Support Laboratory, Cincinnati, Ohio. EPA. 1983 and subsequent EPA - 600/4. 

3 The analytical methods selected for the analysis of critical and noncritical parameters, and the rationale used in their 
selection, are discussed in Section 4.2. 

4 Interim Guidance for Reactive Sulfide. Section 7.3.4.2, SW-846. 

5 Methods for Measuring the Acute Toxicity of Effluents and Receiving Waters to Freshwater and Marine Organisms 
EPA/600/4-90/027F. EPA 1993. 

6 These are field measurements made byTetra Tech. 


26 






and Clear Creek samples. All field measurements were 
made in accordance with standard operating procedures. 

3.4 Site Demonstration Results 

This section presents the results of the C WS demonstration 
conducted from January 1994 to November 1997. Initially, 
aqueous chemistry data for the Burleigh Tunnel mine 
drainage are presented, followed by the demonstration 
results for the two CWS cells (Sections 3.4.1 through 
3.4.3). 

Section 3.4.4 presents data for the receiving stream, Clear 
Creek, and Sections 3.4.5 and 3.4.6 present toxicity 
results. Tables summarizing analytical results for 
the Burleigh Tunnel mine drainage are included in Appendix 
A. An evaluation of demonstration data quality parameters 
for critical analyses is contained in Section 4. 

The data discussed in this section were generally collected 
using demonstration sampling and analysis techniques. 
However, influent and effluent data for much of 1996 were 
collected and analyzed by the CDPHE laboratory 
(Analytica, in Broomfield, Colorado). In addition, data 
was not collected by Tetra Tech or CDPHE for 3 months 
(September through November) in 1996. Tetra Tech 
discontinued CWS sampling at the end of its initial SITE 
contract and the resumption of sampling was slowed by 
contractual delays. 

3.4.1 Burleigh Mine Drainage Chemistry 

The Burleigh Tunnel drains a network of interconnected 
mines on Republican Mountain and Sherman Mountain. 
Unlike many metal mine drainages, the Burleigh Tunnel 
effluent has near-neutral pH and carbonate alkalinity of 
approximately lOOmg/L. 

The mine drainage contains high levels of zinc that 
typically range from 45 to 65 mg/L. However, in May and 
June 1995, a great deal of spring snow and rain and a rapid 
thaw combined to increase the amount of runoff entering 
the mine network drained by the Burleigh Tunnel. At that 
time, flow from the tunnel increased from 45 gpm to more 
than 300 gpm, and zinc concentrations increased from 55 
mg/L (April 12, 1995) to 109 mg/L (August 8, 1995). 

Over the final 2 years of the demonstration, zinc 
concentrations in Burleigh Tunnel mine drainage were 
lower in the winter, dropped again in April or May when 
flow through the mine workings increased, and rapidly 
increased in summer, remaining high throughout the fall. 


During this period, Burleigh Tunnel mine drainage 
zinc concentrations generally remained between 45 and 
84 mg/L, with increases to more than 100 mg/L noted 
during the late summer and fall. Zinc concentrations in 
Burleigh Tunnel mine drainage between September and 
November 1996 are assumed to be similar to zinc 
concentrations measured during the same period in 1995. 
Figure 4 shows zinc concentrations for the Burleigh 
Tunnel mine drainage measured during the demonstration. 

In addition to zinc, cadmium, lead, nickel, and manganese 
are also demonstration metals of interest. Cadmium, lead, 
and nickel readily form sulfides and are expected to be 
removed by the CWS. Manganese does not form a stable 
sulfide but was shown to be removed in a short¬ 
term treatability study conducted prior to the demonstration 
(PRC 1993). Cadmium, lead, and nickel levels were 
generally less than 0.1 mg/L in the Burleigh Tunnel mine 
drainage. After the high flow event in 1995, cadmium 
levels increased to concentrations ranging from 0.11 to 
0.26 mg/L. Lead and nickel levels were generally much 
lower than cadmium and did not increase to the same 
extent after the high flow event. 

Anion concentrations also increased during the 
demonstration. Sulfate concentrations in the Burleigh 
Tunnel drainage ranged from 279 to 652 mg/L and also 
increased after the high flow event. Carbonate (total 
alkalinity) concentrations were measured over a relatively 
narrow range of 82.4 to 125 mg/L. The highest carbonate 
concentrations were measured during a 1-month period 
in June and July 1995, corresponding to the period of 
highest flow from the Burleigh Tunnel. The simultaneous 
increases in zinc, sulfate, carbonate, and calcium without 
an increase in pH suggest these mine drainage constituents 
originate from mineral dissolution. Calcite (CaC03) is 
commonly found in hydrothermal vein deposits in 
association with lead-silver-zinc formations (Correns 1969) 
and is also reported in the Silver Plume mining district. 
The high concentration of both zinc and carbonate at near 
neutral pH suggests the Burleigh Tunnel mine drainage is 
a combination of waters from multiple sources. 

3.4.2 Downflow CI/VS 

The downflow cell was operated for approximately 
2'/ 2 years during the demonstration. Over this period, the 
system removed 60 to 95 percent of the zinc contamination 
from the Burleigh Tunnel mine drainage. 

Figure 4 shows zinc concentrations in the Burleigh Tunnel 
mine drainage (influent), and the effluents of both CWS 


27 



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cells. During the first year of operation, influent zinc 
concentrations ranged from 45 to 63 mg/L (average of 

57.1 mg/L) and the amount of zinc removed by the 
downflow cell ranged from 35 to 54 mg/L (average of 

44.2 mg/L). Zinc removal efficiency during the first year 
averaged 77.4 percent. During the second year, zinc 
levels in mine drainage ranged from 53 to 109 mg/L 
(average of 83 mg/L) and downflow zinc removal ranged 
from 41 to 78 mg/L (average of 58 mg/L). Zinc removal 
efficiency during the second year averaged 70 percent. 
Over the final 6 months this cell operated, influent zinc 
levels ranged from 46 to 84 mg/L, while downflow CWS 
zinc removal ranged from 31 to 78 mg/L. In general, 
greatest zinc removal corresponded to times with the 
highest influent zinc concentrations, and the lowest zinc 
removal was observed during periods of lesser zinc in the 
mine drainage suggesting metal removal was effected by 
a physical process. 

Although present only in low levels in the influent water, 
cadmium, lead, and nickel were removed to a great extent 
by the downflow CWS treatment. Influent cadmium 
concentrations ranged from 0.071 to 0.10 mg/L, while 
effluent levels ranged from 0.0007 to 0.003 mg/L during 
the first year. During the second year, cadmium 
concentrations increased in the influent, ranging from 
0.057 to 0.26 mg/L, and downflow effluent levels ranged 
from 0.0001 to 0.007 mg/L with few detections. Figure 5 
shows cadmium concentrations for the influent and both 
effluents during the first 2 years of the demonstration 
Substantial cadmium removal continued over the final 
6 months by the downflow cell, with the exception of the 
April 1996 sample. 

Samples were not regularly analyzed for lead or nickel 
during the demonstration. Figure 6 shows lead 
concentrations for the influent and both effluents during 
the first 2 years of the demonstration. During the first 
year, influent lead concentrations ranged from 0.013 to 
0.020 mg/L, while downflow effluent concentrations 
ranged from 0.00065 to 0.0054 mg/L. Throughout the 
remainder of 1995, influent levels of lead increased 
slightly while effluent levels remained very low with few 
detections. 

Nickel was also removed by the downflow cell; however, 
the extent of removal declined when influent nickel 
concentrations increased after the high flow event. 
Nickel levels in the influent ranged from 0.033 to 0.68 mg/ 
L, and downflow effluent ranged from 0.0073 to 0.020 
mg/Lin the first year. Throughout the remainder of 1995, 


influent nickel levels ranged from 0.045 to 0.093 mg/L, 
and downflow effluent levels ranged from 0.014 to 
0.040 mg/L. 

Manganese concentrations in the mine drainage were 
initially between 1 to 2 mg/L. Manganese removal by the 
downflow CWS was low during the demonstration. Figure 
7 shows manganese concentrations for the influent and 
both effluents. 

The extended residence time of the influent within the 
downflow cell substrate caused by low flow rates may be 
one reason the downflow CWS was effective in removing 
metals from the mine drainage. Both wetland cells were 
designed to treat 7 gpm; however, the permeability of the 
downflow cell declined during the first year of operation, 
and flow through the cell dropped to 4 gpm particularly 
during the summer months. Although attempts were 
made to increase its permeability by fluffing the substrate 
with compressed air, these procedures resulted in only 
temporary improvements. Flow through the downflow 
cell improved during winter months when the substrate 
froze and contracted from the liner allowing the influent to 
flow down the sides of the interior cell. Flow through the 
downflow cell averaged 6.5 gpm during the first year; 5.8 
gpm in the second year; and 6 gpm over the final 6 months 
of operation. 

Analytical results for the downflow substrate (Table 4) 
showed a substantial increase in zinc levels over the 
period of the demonstration. Substrate zinc levels ranged 
from a low of 59.7 milligrams per kilogram (mg/kg) to a 
high of 5,630 mg/kg. Substrate samples were generally 
collected from between 1 to 2 feet below the surface of 
the CWS. Downflow substrate samples contained little 
visible evidence of sulfate reduction and low concentrations 
of AVS. Sulfate-reducing bacteria counts showed much 
variability (Figure 8). 

After the first 6 months of operation, the downflow cell 
was removing more zinc from the mine drainage compared 
with the upflow cell. However, the reason for the greater 
removal was likely the higher residence time of the mine 
drainage within the downflow wetland. The increasing 
residence time was a function of mine drainage flow 
through the cell, that was generally lower in the summer 
compared to winter. A reduction of flow from 7 to 5 gpm 
increases residence time by 19 hours nearly a 40 percent 
increase. The loss of permeability is believed to be related 
to the loss of permeability in the downflow cell resulting 
from biological surface growth, chemical precipitation of 


29 




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Table 4. Average Downflow CWS Substrate Results 



Cadmium 

(mg/kg) 

Lead 

(mg/kg) 

Nickel 

(mg/kg) 

Zinc 

(mg/kg) 

Acid Volatile 
Sulfides 
(mg/kg) 

Sulfate- 

Reducing 

Bacteria 

(count) 

Ortho¬ 

phosphate 

(mg/kg) 

0-6 months 

2.7 

18 

3.1 

1,100 

180 

8.5 x 10 4 

34 

6-12 months 

8.0 

31 

6.1 

3,400 

120 

1.1 x 10 5 

12 

12-18 months 

23 

74 

7.0 

5,200 

460 

3.3 x 10 5 

2.6 


Notes: 

mg/kg Milligram per kilogram 

Average Arithmetic Mean 

Substrate samples collected from 1-2 feet below wetland surface 


zinc compounds, microbial breakdown of the substrate to 
finer particulates, and the settling of these particles into 
substrate pore spaces. The increase of flow during winter 
is believed to result from freezing of the wetland substrate 
at the edge of the cell causing the substrate to contract 
from the liner. The contraction allowed ponded water at 
the surface of the wetland to flow between the frozen 
substrate and liner to the base of the cell forming a 
preferential pathway. 

Loading is the amount of metals retained by the wetland 
over time. It is a function of the flowrate through the 
wetland, the concentration of metals in the mine drainage, 
and the removal efficiency of the treatment. For this 
discussion, monthly loading of each wetland was calculated 
from measured flow rates and simultaneously collected 
samples of the mine drainage and the wetland effluent. 
Figure 9 shows the monthly zinc loading to the downflow 
CWS over the demonstration. The graph indicates that 
loading was initially high (maximum of 60 kg/month) but 
dropped as the downflow cell flow rate declined in the Fall 
of 1994. In winter, loading also increased as flow 
improved. The greatest loading to the downflow CWS 
occurred during the high flow event in the late spring and 
early summer of 1995. After the high flow event, loading 
in this cell declined dramatically and eventually dropped to 
less than 5 kg/month in May 1996. 

The primary metal removal mechanism active in this cell 
did not appear to be sulfate reduction. Substrate analyses 
indicate a significant portion of the zinc removal in this 
CWS occurred in the upper 1 to 2 feet of substrate, where 
few A VS or sulfate-reducing bacteria were found. Pockets 
of sulfide-rich substrate were observed in this CWS cell 
at depths of 3 to 4 feet below the wetland surface, 


suggesting some sulfate reduction contributes to metal 
removal in this wetland. Aqueous geochemical modeling 
of the mine drainage suggests gypsum is oversaturated; 
however, visual observations of Burleigh Tunnel mine 
drainage precipitate and historical mine reports suggest 
the material is a zinc carbonate, probably smithsonite or 
hydrozincite. 

The following can be concluded from the evaluation of the 
downflow CWS: 

• As tested, the downflow CWS did not retain sufficient 
permeability to be considered a reasonable long¬ 
term treatment option. 

• Chemical precipitation (suspected to be mineral 
carbonate accumulations) may have been the primary 
metal removal process in this CWS treating Burleigh 
Tunnel mine drainage. 

• A 2-foot substrate depth should be adequate, as 
most metal removal occurred at between 1 to 2 feet 
below the wetland surface. A thinner substrate 
should decrease the flow resistence of the downflow 
CWS and increase the effectiveness of the system. 

• A 2-foot downflow CWS may be a good pretreatment 
for an upflow CWS treating the Burleigh Tunnel 
mine drainage allowing some physical precipitation 
of the zinc. 

The concentration of orthophosphate in the substrate also 
decreased after the high flow event in 1995. The high 
orthophosphate concentration, measured at the beginning 
of the demonstration, was 114 mg/kg; the low, 1 to 2 mg/ 
kg, was measured in August 1995. 


34 






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>6-d9S 


1 ^ 6 —i n r 


t6-^DlA| 


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(L> 

O 

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35 


Figure 9. Monthly zinc loading, downflow CWS. 




3.4.3 UpflowCWS 


The upflow cell was demonstrated for nearly 4 years and, 
during this period, removed zinc and other metals initially 
by adsorption, later by sulfate reduction, and eventually by 
chemical precipitation (presumed). The adsorption period; 
appeared to last roughly 4 to 5 months as indicated by 
manganese removal. After the adsorption phase, sulfate 
reduction appeared to be the primary metal removal 
process; however, oxidation/reduction (ORP) 
measurements suggested the activity of the sulfate- 
reducing bacteria appeared to drop in late fall and through 
the winter of 1994. Counts of sulfate-reducing bacteria 
declined coincidentally with the decline in ORP. The drop 
may have been caused by lower winter temperatures, or 
an increase in flow through the cell that occurred in 
September through October 1994, or may result from the 
use of all the most easily metabolized materials in the 
compost substrate by the bacteria. During this period, the 
concentration of zinc in the upflow effluent increased 
from3.2 mg/L(October 12,1994)to 18 mg/L (March 15, 
1995). 

By May 1995, zinc levels were approaching levels that are 
inhibitory to sulfate-reducing bacteria at the observed 
area loading of 250 square feet per gallon. During May 
and June of that year, the high flow event exposed the 
wetland sulfate-reducing bacteria to elevated levels of 
zinc, and the high influent flow probably created aerobic 
conditions within the cell. The periodic high zinc 
concentrations observed in influent waters during the 
summer and fall of 1996 and 1997 likely prevented the 
sulfate-reducing bacteria from reestablishing activity to 
previous levels. The flow was halted to the upflow cell in 
the summer of 1997 for approximately one month for 
repairs. At that time, much of the water was removed 
from the cell, allowing wetland sulfate-reducing bacteria 
an opportunity to become reestablished. 

However, there was no indication that the bacteria became 
re-established during the final 4 to 5 months of the 
demonstration. One of the repairs involved plugging a 
short section of the influent piping in the upflow cell. 
Visible observation of this influent pipe noted a black 
coating on the inside of approximately 1/16 inch and 
accumulations of black precipitate nearly filling the holes 
in the perforated pipe. Overlying the black material in the 
piping was a layer of cream colored to yellow material up 
to 1/8 of an inch thick. 

Analytical results for influent and effluent samples from 
the upflow system showed that zinc was nearly completely 


removed by this system during the first 8 months of the 
demonstration (Figure 4). After this period, zinc 
concentrations in the upflow effluent gradually increased 
from 1.4 mg/L (September 19, 1994) to 18.5 mg/L in the 
spring of 1995 corresponding to zinc removal efficiencies 
of 97.6 and 66.8, respectively. In May and June 1995, high 
flow from the Burleigh Tunnel increased flow through the 
upflow cell to 20 gpm and zinc concentrations nearly 
doubled. Over the next 6 months, as flow decreased from 
the tunnel, influent zinc concentrations rose to a high of 
109 mg/L. From May to November 1995, effluent zinc 
levels increased from 26.7 to 73.6 mg/L. The amount of 
zinc removed by the upflow cell averaged 41 mg/L (49.3 
percent) during the second year. 

During the third year of operation, zinc levels in the 
influent ranged from 56 to 84 mg/L; however, data were 
not collected between September and November 1996. 
Zinc concentrations in the upflow effluent over the third 
year ranged from 30 to 49 mg/L with an average removal 
of 30 mg/L (39.6 percent). In the final year of operation, 
zinc influent concentrations ranged from 42 to 104 mg/L 
and effluent levels ranged from 15 to 60 mg/L with an 
average removal efficiency of 65.1 percent. Effluent 
levels were greater in the May 28, 1997 sample (60 mg/ 
L) compared to the influent sample (56 mg/L). Over the 
final 6 months, the upflow cell removed greater amounts 
of zinc as flow through the cell decreased. Flow through 
the upflow cell at this time ranged from 2 to 5 gpm. 

Cadmium removal by the upflow cell followed a pattern 
similar to zinc removal (Figure 5). Initially, cadmium was 
removed to nondetect levels; however, cadmium 
concentrations increased two and a half times after the 
high flow event. After this period, cadmium removal 
remained high for 4 months but declined in the latter part 
of 1995 and remained low through 1996 and 1997. 

Lead (Figure 6) and nickel were also removed to lower 
concentrations by the upflow CWS. Influent lead and 
nickel concentrations were approximately 0.015 mg/L 
and 0.043 mg/L, respectively. During the first year, lead 
was removed to nondetect levels and nickel effluent 
concentrations ranged from 0.0005 to 0.019 mg/L. Unlike 
zinc and cadmium, lead and nickel concentrations did not 
increase significantly after the high flow event; however, 
the removal ofboth decreased somewhat until flow values 
through the cell declined in the final months of the 
demonstration. 

Manganese was initially present in the mine drainage at 
concentrations ranging from 1 to 3 mg/L. Manganese 


36 





was removed by the upflow cell for the first 4 months of 
operation but was not removed throughout the remainder 
of the demonstration. 

Analytical results for the upflow substrate showed an 
increase in zinc levels over the period of the demonstration. 
Table 5 summarizes mean annual results for selected 
analysis from upflow cell substrate samples collected 
during the demonstration. Zinc levels ranged from a low 
of 40 mg/kg to a high of4,800 mg/kg. The zinc content is 
expected to be higher in the removal zone of the upflow 
cell (deeper in the substrate of the cell). In general, upflow 
substrate samples were collected approximately 2 feet 
below the wetland surface, above the removal zone. 
Counts of sulfate-reducing bacteria in the upflow cell 
were generally very high between April 1994, through 
July 1995. However, counts were 1 to 2 orders of 
magnitude lower in upflow cell samples collected in 
April 1996 through September 1997. The final substrate 
sample analyzed for sulfate-reducing bacteria 
contained approximately 250,000 CFU/gram substrate. 
Figure 10 shows the results of sulfate-reducing bacteria 
counts conducted on upflow cell substrate samples 
collected during the demonstration. 

The change from strongly reducing to slightly reducing 
conditions in the fall of 1994 may have made previously 
removed metal sulfides less stable within the wetland 
substrate. Substrate observations in the summer of 1997 
indicated there were fewer sulfides present compared to 
substrate samples collected in 1994 and 1995. If half of 
the zinc removed in the first year of operation were 
released over the subsequent 2 years, the resulting zinc 


increase in the effluent would have been 33 mg/L. The 
higher zinc concentration measured in the May 28, 1997 
effluent sample compared to the corresponding influent 
sample suggests some previously removed zinc was 
released. 

Between March and December 1994, metals loading to 
the upflow CWS ranged from 53 to 97 kg/month but 
dropped to 26 kg/month in February 1995. This drop in 
loading corresponded with the increase of zinc in the 
effluent, an increase in ORP, and a decrease in flow rate 
through the cell. Flow through the cell increased in March 
and April 1995, leading to higher loading. The maximum 
loading to the upflow CWS (107 kg/month) occurred in 
May 1995 during the high flow event. Throughout the 
remainder of the demonstration, loading to this cell declined 
as the zinc removal efficiency decreased to 40 to 50 
percent; eventually, flow through the cell ended in 1997. 
Figure 11 shows zinc loading to the upflow CWS over the 
demonstration. 

The effect of the high flow event on the performance of 
the upflow CWS reveals the major shortcoming of passive 
systems, the inability to adapt to rapidly changing conditions. 
In this demonstration, the upflow CWS could not adjust to 
the increased influx of zinc or the change in environmental 
conditions. 

As several constructed wetlands have successfully treated 
mine drainage with much higher concentrations of zinc, it 
may be concluded that the bacteria are somehow able to 
protect themselves from the high metals concentration. If 
this mechanism is sulfate reduction, the rate of sulfate 


Table 5. Average Upflow CWS Substrate Results 


Acid Volatile Sulfate- Ortho- 



Cadmium 

(mg/kg) 

Lead 

(mg/kg) 

Nickel 

(mg/kg) 

Zinc 

(mg/kg) 

Sulfides 

(mg/kg) 

Reducing 
Bacteria (count) 

phosphate 

(mg/kg) 

Year 1 

0.17 

9.9 

1.9 

40 

210 

7.2 x 10 6 

55 

Year 2 

0.18 

13 

2.0 

71 

460 

3.2 x 10 6 

54 

Year 3 

5.0 

40.0 

4.1 

1,500 

1,300 

2.2 x 10 5 

6.3 

Year 4 

9.6 

NR 

6.2 

4,800 

1,000 

6.2 x 10 4 

6.9 


Notes: 

mg/kg Milligram per kilogram 

NR Not Reported 

Average =Arithmetic Mean 

Substrate samples collected from 1-2 feet below wetland surface 


37 





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39 


Figure 11. Monthly zinc loading, upflow CWS. 






reduction must be great enough to reduce zinc 
concentrations in the substrate to below inhibitory levels. 
This hypothesis suggests that the effectiveness of an 
anaerobic compost C WS is a function of the rate of sulfate 
reduction, residence time of the mine drainage in the 
wetland substrate, and the concentration of zinc (or other 
inhibitory metals) in the mine drainage. Low temperature 
is also a factor that will affect the activity of sulfate- 
reducing bacteria in the wetland. 

The following can be concluded from the evaluation of the 
upflowcell: 

• The upflow CWS is effective in removing many 
metal contaminants from mine drainage; however, 
the CWS may have difficulty recovering from rapidly 
increasing metals loading conditions. Reinnoculation 
and incubation of sulfate-reducing bacteria may 
improve recovery of these systems. 

• Control of mine drainage flow to the constructed 
wetland is critical to ensure that residence time and 
operational conditions are maintained. 

• The operational lifetime of an upflow CWS (with a 
compost substrate depth of 4 feet) is roughly 4 to 
5 years. 

• The upflow cell had superior hydraulic performance 
throughout most of the demonstration. 

• Winter freezing can be prevented by covering the 
wetland surface with hay or blankets used in curing 
concrete. 

• Piping cleanouts should allow all piping networks to 
be easily cleaned. 

3.4.4 Clear Creek 

The untreated Burleigh Tunnel mine drainage and the 
effluents of both CWS cells discharge to Clear Creek. To 
assess the impact of treatment on the receiving stream, 
upstream and downstream samples collected from Clear 
Creek were also analyzed for total metals and aquatic 
toxicity. The metals results indicated that although the 
wetlands may be removing metals from the mine drainage, 
the demonstration-scale C WS treated only a small portion 
of the total discharge from the Burleigh Tunnel, not 
enough to show a measurable decrease in the metals 
content of the stream. The demonstration-scale CWS 
treated approximately 30 percent of the total flow from 
the Burleigh Tunnel, and during high flow treated only 
about 5 percent of the flow. A full-scale system could 
show a more significant decrease in the metals content of 
Clear Creek downstream of the system. 


The stream results for upstream versus downstream 
samples are presented in Tables 6 and 7. The results show 
that Burleigh Tunnel mine drainage is a significant source 
of zinc to Clear Creek. However, CDPHE reports there 
are also additional nonpoint sources of zinc-contaminated 
water received by the creek. 

3.4.5 Toxicity Testing Results 

Constructed wetland treatment is a complex 
biogeochemical process involving adsorption, chemical 
precipitation, and microbial interactions with contaminants. 
The primary metal removal mechanisms in the CWS are 
chemical precipitation and microbial sulfate reduction; 
however, treatment may also complex metal contaminants, 
making them unavailable to receptor organisms. Thus, 
aquatic toxicity analyses were conducted by the EPA 
National Exposure Research Laboratory - Aquatic Toxicity 
during the demonstration to evaluate the reduction in 
toxicity resulting from CWS treatment. Two test organisms 
were used in the toxicity testing: water fleas (Ceriodaphnia 
dubia) and fathead minnows (Pimephales promelas). A 
total of eight rounds of aquatic toxicity testing were 
conducted during the demonstration. Initially, toxicity 
samples were collected and analyzed every 3 to 4 months 
until late 1995, when demonstration activities were 
temporarily suspended. When demonstration monitoring 
resumed, toxicity testing was conducted every 4 to 
6 months. In 1997, a microbial toxicity test was conducted 
on wetland sulfate-reducing bacteria with Burleigh Tunnel 
mine drainage. The results of the microbial toxicity test 
are presented in Section 3.4.6. 

Aquatic toxicity testing results correlated well with 
increasing zinc concentrations observed in the effluents of 
the treatment cells during the first 2 years of the 
demonstration. Results of testing conducted during the 
first 8 months of the demonstration indicate the effluents 
from both cells were not toxic to either the C. dubia or the 
P. promelas. The Burleigh Tunnel mine drainage was 
toxic to both test organisms at low concentration (dilution) 
throughout the demonstration. Table 8 provides influent 
and effluent concentrations resulting in the death of 
50 percent of the test organisms (LC50) in each round of 
testing. As zinc concentrations increased in the effluents 
of both cells through 1995, so did the toxicity to the test 
organisms. 

The first test conducted that year (February 1995) indicated 
that effluent from the upflow cell had become toxic to 
C. dubia at a concentration of 8.4 percent. The high 
runoff event that occurred in the spring of 1995 and 


40 



Table 6. Clear Creek Upstream 



Cadmium 

(mg/L) 

Lead 

(mg/L) 

Nickel 

(mg/L) 

Zinc 

(mg/L) 

pH 

Conductivity 

(MS) 

Temperature 

( # C) 

Average 

0.0022 

0.0034 

0.0047 

0.126 

7.8 

155.7 

5.4 

Maximum 

0.0094 

0.013 

0.015 

0.56 

8.1 

167.5 

9.7 

Minimum 

0.0 

0.0 

0.0 

0.11 

7.6 

144.0 

0.9 


Notes: 

°C Degrees Celsius 
mg/L Milligrams per liter 

pS MicroSiemens 

ND Not Detected 

pH Standard units 

Average sArithmetic Mean 


Table 7. Clear Creek Downstream 



Cadmium 

(mg/L) 

Lead 

(mg/L) 

Nickel 

(mg/L) 

Zinc 

(mg/L) 

pH 

Conductivity 

(MS) 

Temperature 

(°C) 

Average 

0.00075 

0.0013 

0.0068 

0.512 

7.6 

132.8 

4.3 

Maximum 

0.0017 

0.0024 

0.026 

0.56 

8.1 

173.3 

9.7 

Minimum 

ND 

ND 

ND 

0.14 

6.5 

80.0 

.. 


Notes: 


°C Degrees Celsius 
mg/L Milligrams per liter 

pS MicroSiemens 

ND Not Detected 

pH Standard units 

Average EArithmetic Mean 


associated increases in flow through the CWS cells and 
elevated zinc concentrations resulted in higher zinc levels 
in the CWS effluents. At that time, the effluent from both 
cells became toxic to the test organisms. The upflow cell 
effluent was toxic to C. dubia at a concentration of 
0.1 percent and to P. promelas at concentrations ranging 
from 1.2 to 2.3 percent. The downflow cell effluent was 
toxic to C. dubia at concentrations ranging from 0.31 to 
0.51 percent and to P. promelas at concentrations ranging 
from 2.6 to 30 percent. 

Over the final 2 years of the demonstration, the upflow cell 
effluent continued to be toxic to C. dubia at concentrations 
below 1 percent and to P. promelas at a concentration of 
14 percent. Toxicity samples were not collected from the 
downflow cell: operation of this cell was discontinued in 
September 1996. 


Demonstration toxicity testing results indicate that the 
ability of the wetlands to reduce toxicity to aquatic 
organisms gradually declined over the first 2 years. In 
addition, the high flow event in 1995 had a significant 
impact on zinc and toxicity removal by the upflow cell over 
the final 2 years of the demonstration. 

Water samples for toxicity testing were collected from 
Clear Creek above and below the CWS discharge three 
times during the demonstration. As mentioned, the 
constructed wetlands treated only 30 percent of the mine 
drainage; thus, the impact of treatment on the receiving 
stream was minor. One set of samples contained higher 
toxicity in the upstream sample while samples collected 
after June 1995 indicated that there was no acute toxicity 
in the upstream samples but that addition of the mine 
drainage to the stream resulted in an increase in toxicity. 


41 







Table 8. CWS Demonstration Toxicity (LC J0 ) Results 


Indicator Species 

Date 

Collected 

Influent 

Upflow 

Effluent 

Downflow 

Effluent 

Clear Creek 
Upstream 

Clear Creek 
Downstream 

Fathead Minnows 

08/24/94 

1.1 

No toxicity 

NA 2 

No toxicity 

No toxicity 

(Pimephalus 

promelas) 

09/19/94 

0.73 

No toxicity 

No toxicity 




02/22/95 

1.6 

No toxicity 

No toxicity 




06/12/95 

1.0 

2.3 

2.6 

No toxicity 

No toxicity 


09/05/95 

0.62 

1.2 

30 




12/10/96 

0.62 

1.6 

NA 




06/24/97 

0.69 

24 

NA 

No toxicity 

No toxicity 


10/29/97 

1.4 

14 

NA 




10/29/97 1 


11 




Water Fleas 

08/24/94 

0.46 

No toxicity 

NA 

No toxicity 

No toxicity 

(Ceriodaphnia dubia) 

09/19/94 

0.31 

No toxicity 

No toxicity 




02/22/95 1 

1.0 

8.4 

No toxicity 




02/22/95 



No toxicity 




06/12/95 

0.10 

0.43 

0.51 

No toxicity 

No toxicity 


12/10/96 

0.09 

0.22 

NA 




06/24/97 

0.43 

0.41 

NA 

No toxicity 

No toxicity 


09/05/95 

0.10 

<0.19 

0.31 




10/29/97 

0.15 

0.13 

NA 




10/29/97 1 


0.19 

NA 




Notes: 

1 Duplicate Sample 

2 NA-Not analyzed 


3.4.6 Microbial Toxicity Testing 

Microbial toxicity testing was undertaken when repairs to 
the upflow cell indicated that there were few metal 
sulfides in the wetland substrate compared with 
observations conducted in previous years. The lack 
of metal sulfide deposits in the substrate suggested 
that the sulfate-reducing bacteria were not actively 
producing sulfide. Thus, Burleigh Tunnel mine drainage 
was tested at the Colorado School of Mines for toxicity to 
sulfate-reducing bacteria isolated from the upflow cell. 

The tests indicated that the mine drainage is inhibitory to 
sulfate-reducing bacteria at low concentrations (dilution) 
corresponding to a zinc concentration of 17.5 mg/L. 
In addition, zinc sulfate (ZnS04-7 H20) was used to 
show that the zinc was the toxic constituent (positive 


control) in the mine drainage. The zinc sulfate was also 
toxic to the sulfate-reducing bacteria at a similar zinc 
concentration (18.8 mg/L). The concentration of zinc in 
the Burleigh Tunnel mine drainage typically exceeds the 
inhibitory level measured in this study. A similar study 
conducted using Desulfovibrio desulfricans also found a 
zinc concentration of 13 mg/L resulted in inhibition to the 
bacteria. (Paulson and others 1997). 

Evidence that sulfate reduction was important to the 
removal of zinc in the upflow CWS include the large 
population of sulfate-reducing bacteria observed when 
zinc removal was also high (first year of demonstration), 
the accumulation of A VS, primarily zinc sulfide, in the 
substrate of this cell, and the decline of sulfate-reducing 
bacteria populations after the high flow event that 
corresponded with lower zinc removal by the upflow cell. 


42 





Visible observations of the up flow cell substrate observed 
blackening of the substrate during the first year of operation 
suggesting metal sulfides were accumulating, however, 
observations of wetland substrate conducted three years 
later, showed little blackening of the substrate. These 
results suggest sulfate-reduction was not as an important 
metal removal mechanism and was occurring to a much 
lesser extent during the latter portion of the demonstration. 
These observations also suggest that previously formed 
metal sulfides are not stable when environmental conditions 
within the wetland changes. 

3.5 Attainment of Demonstration 
Objectives 

This section discusses the results of the C WS demonstration 
in regard to the attainment of primary and secondary 
demonstration objectives. In addition, metal removal 
mechanisms, some of the causes for poor performance, 
and substrate lifetimes are discussed for each cell. 

The results of the demonstration were able to achieve 
many but not all of the primary objectives outlined in 
Section 3.3. The first primary objective was the 
measurement of wetland effectiveness with respect to 
cell flow configuration and seasonal variation. This 
primary objective was achieved in part. The demonstration 
zinc results indicate zinc removal is greater with an up flow 
configured wetland; however, the technology as tested is 
not capable of meeting low metal discharge requirements 
for extended periods. 

The better zinc removal and flow of the mine drainage 
through the upflow CWS compared to the downflow 
CWS indicate the upflow configuration is superior. 
Unfortunately, it was not possible during this demonstration 
to determine the effect of season variation on the 
performance of the upflow CWS. The downflow CWS 
actually performed better during the winter. The reason 
for the improved winter performance is discussed in 
Section 3.4.2. 

The second primary objective was to determine the 
toxicity of the Burleigh Tunnel mine drainage. This 
primary objective was achieved. The Burleigh Tunnel 
mine drainage is toxic to both the C. dubia and P. 
promelas. Measured LC50 values for the P. promelas 
(fathead minnows) ranged from 0.62 to 1.6 percent (mine 
drainage) and for the C. dubia (water fleas) ranged from 
0.10 to 1.0 percent. 


The third primary objective was the characterization of 
toxicity reduction resulting from CWS treatment. This 
primary objective was also achieved. The demonstration 
toxicity results indicate the ability of the wetlands to 
reduce toxicity to aquatic organisms declined over the first 
two years of operation. Further, the high flow event had 
a significant impact on toxicity removal in both wetland 
cells. 

The final primary objective was to estimate the toxicity 
reduction to the mine drainage receiving stream (Clear 
Creek). This primary objective was not achieved as none 
of the demonstration stream samples were toxic to either 
test organism. 

The most significant primary objective not achieved is the 
inability to detemiine the seasonal variability of the upflow 
CWS. During winter, constructed wetlands located in 
cold climates may be less effective as a result of lower 
microbial activity. This may require pretreatment of the 
mine drainage during winter, oversizing the CWS or 
retaining a portion of the flow until warmer conditions 
return. 

The first secondary objective of the demonstration was 
to estimate the lifetime of the substrate material. The 
lifetime of substrate material is estimated to be 4 to 5 
years. The estimate is based on the breakdown of the 
substrate material resulting in settling and compaction of 
the substrate that leads to flow restrictions. In addition, 
demonstration substrate data for nutrients indicate 
elements such as phosphate (orthophosphate) have been 
depleted in the substrate by this time. If low discharge 
limits must be met then demonstration results suggest the 
substrate lifetime is approximately one year (taking into 
account the demonstration starting time and freezing of 
the upflow cell during the first year). However, in this 
situation it would likely be more cost effective to pretreat 
the mine drainage or amend it with an electron donor such 
as ethanol to extend the lifetime of the substrate material. 

The second, noncritical or secondary objective was to 
estimate metal removal by sulfate reducing bacterial. 
This evaluation was expected to be qualitative as the 
bacteria counts and acid-volatile sulfide analyses are not 
highly precise and the metal removal may not be uniform 
throughout the treatment cells. As discussed in Section 
3.4.2, the downflow cell data did not indicate the primary 
metal removal mechanism to be sulfate reduction. Section 
3.4.3 discusses the upflow cell results for sulfate-reducing 
bacteria removal of metals. Data indicated an initial high 


43 



rate of removal with a longer term reduction in this 
mechanism of metals removal. 

The third noncritical, secondary objective was to evaluate 
the impact of the systems effluent on Clear Creek. These 
data are discussed in Section 3.4.4, and indicate that 
although the treatment was effective in removing metals 
from the Burleigh Tunnel drainage, the relatively small 
portion of the discharge being treated did not produce a 
measureable decrease in the metals content of Clear 
Creek. 

The fourth and final noncritical objective was to evaluate 
capital operating costs for the CWS. Section 5.0 of this 
report provides a detailed economic analysis and 
successfully provides data useful for estimating costs for 
application of this technology at other sites. 

3.6 Design Effectiveness 

The following sections discuss the effectiveness of the 
upflow and downflow CWS tested during the Burleigh 
Tunnel demonstration. The basic design of each wetland 
cell is discussed in Section 1.3.2 of this report. This 
discussion focuses on general design parameters and 
factors that affected each cell. 

The basic design of the CWS demonstration system 
consisted of a dam inside the Burleigh Tunnel, piping from 
the dam to the influent weir, the two wetland cells, an 
effluent weir, and a bypass pipe. The dam collected the 
mine drainage and provided adequate hydraulic head to 
drive the mine drainage through the upflow cell. The 
influent weir partitioned the mine drainage to the CWS 
cells and channeled the excess water to the bypass piping. 
From the influent weir, the mine drainage was channeled 
to a ball valve that separated flow to the CWS cells. Water 
collected from the cells was piped to the effluent weir and 
was discharged to Clear Creek. The purpose of the 
effluent weir was to regulate flow through the wetland 
cells. 

Construction materials associated with this design were 
generally inexpensive, readily available, and easily 
transported to remote areas. Installation techniques were 
also straightforward. 

The major drawbacks of this design observed during the 
demonstration centered on the flow control valves and 
the inability of the effluent weir to regulate flow through 
the cells. Because flow through the cells could not be 
controlled with the effluent weir, flow through the cells 


was regulated at the influent weir and control valve. 
Unfortunately, this design meant that any adjustment in 
flow to one cell affected flow to the other cell. Future 
systems should use easily controlled flow structures such 
as weirs to regulate flow to both cells independently. 

In addition, the capacity of the initial 4-inch bypass line 
was insufficient to accommodate the large water volume 
during spring runoff. Eventually, a 6-inch bypass line was 
installed. Piping connecting the influent control structure 
and the cells should be direct and accessible for routine 
cleanout. 

A drawback associated with the use of compost substrates 
is the high concentration of nitrate in the effluent water 
during startup. During this demonstration, no attempt was 
made to remove the nitrate from the water prior to 
discharge. In a similar wetland evaluation, startup effluents 
were applied to surface soils. Alternatively, the startup 
effluent could be stored on site in a pond or tank and fed 
back into the CWS. 

3.6.1 Downflow Cell 

The downflow cell consisted of 4 feet of a compost (95 to 
96 percent) and hay (4 to 5 percent) substrate. The mine 
drainage flowed from the top to a PVC piping collection 
network at the base of the cell. The influent and effluent 
distribution networks were staggered within the cell to 
minimize short-circuiting of the mine drainage in the 
substrate. 

The design of the downflow cell is discussed in 
Section 1.3.2; Figure 2 shows a cross section of the 
anaerobic CWS in an upflow configuration. The downflow 
configuration is only a reversal of the influent and effluent 
flows, not the construction of the cell. 

For the most part, the materials used in the construction of 
the cells-HDPE liner, geonets, and PVC piping were 
acceptable. However, the geofabric was found to fill with 
fine material and lose permeability over the 2'/ 2 -year 
demonstration. In addition, the cell piping networks did not 
include cleanouts. Cleanouts should be included in future 
CWS designs. Finally, the influent piping network did not 
evenly distribute the mine drainage in this cell. An 
additional row of perforated piping in this cell would more 
evenly distribute the mine drainage. 

The cell was designed to treat 7 gpm. However, during 
the demonstration, the downflow cell became less 
permeable. The permeability loss is believed to be related 


44 



to precipitation of metal oxides, hydroxides, and carbonates, 
settling of fme materials in the cell, and compaction of the 
substrate material. In winter months, flow through the 
downflow cell improved; presumably, the contraction of 
frozen substrate allowed water to flow between the liner 
and the substrate. However, this short circuiting did not 
substantially affect metal removal by the cell. 

In an attempt to restore flow through the downflow cell, 
air was injected into the substrate to fluff the material. 
Although this technique improved flow, the effect was 
typically short lived. The results of this demonstration 
indicate that substrates with high concentrations of compost 
will not retain permeability in a downflow configuration 
and are not recommended. However, some recent 
downflow wetlands have used substrate mixtures of 50 
percent limestone with sawdust and compost to improve 
hydraulic characteristics. 

3.6.2 Upflow Cell 

The design of the upflow C WS is identical to the downflow 
cell except that the mine drainage is channeled up though 
the compost substrate. Figure 2 shows a cross section of 
the demonstration anaerobic compost CWS. The design 
of the demonstration wetlands is discussed in Section 
1.3.2. 

In general, the upflow cell retained permeability throughout 
the demonstration. However, some hydraulic restriction 
developed during the later half of the demonstration 
resulting in a preferential flow pathway. In addition, gas 
buildup produced by fermenative bacteria within the 
upflow cell may have restricted flow to the effluent lines 
in the wetland during the last year of the demonstration. 
Gas was released from the cell by periodically puncturing 
the upper geofabric with a pitch folk. Replacing the 
geofabric with a fme mesh geonet could eliminate gas 
buildup. Also, the decline of sulfate-reducing bacteria and 
apparent increases in the population of fermentative 
bacteria likely exacerbated the problem. 

The upflow cell was prone to freezing during winter. 
During startup, the dike within the Burleigh Tunnel gave 
way, stopping flow to the upflow cell. Flow was restored 
by thawing the ice around the effluent line with a steam 
cleaner and water tank heater. The following winter, hay 
bales were placed over the substrate followed by insulated 
blankets (identical to insulated blankets used for curing 
concrete), and the system was operational throughout the 
winter. However, the straw bales became saturated with 
water and the combined weight compressed the substrate 


so that all flow ceased through the cell. Flow through 
the cell was restored once the hay bales were removed. 
During year three, the insulated blankets were used alone 
to insulate the cell and there were no interruptions in flow 
during this period. In the final year, the ponded water in 
the upflow cell was allowed to freeze and did so to a depth 
of approximately 6 inches. There were no interruptions in 
flow during that winter. 

Residence time is an important factor in anaerobic 
constructed wetlands that use sulfate-reducing bacteria. 
Decreasing residence times may overload the wetland, 
exposing the bacteria to inhibitory concentrations of zinc. 
Based on the size of the wetlands and substrate water 
volumes (percent moisture results of 50 percent) the 
calculated residence time for a flow rate of 7 gpm is 
48 hours, and 67 hours at a flow rate of 5 gpm. Verification 
of residence times was one of the more difficult 
measurements undertaken during the demonstration. Both 
a chloride tracer (treatability study) and an organic dye 
test (demonstration) were unsuccessful in measuring 
residence time. The chloride could not be readily measured 
as background levels of dissolved salts was somewhat 
high during the treatability study and the organic dye likely 
absorbed to the wetland substrate during this demonstration 
test. 

During the final year of the demonstration, flow through 
the upflow cell began to short circuit in an area adjacent 
to the southeastern bermed sidewall. An excavation was 
made into the wetland to the influent line feeding this 
section of the cell and the line was capped. Dewatering 
the excavation was somewhat difficult and would have 
been aided by a sump within the cell. Inspection of the 
influent line found precipitates coating the piping walls and 
in the piping perforations. The amount of material in the 
perforations and the pressure on the piping against the 
geofabric would have caused a notable restriction in flow. 
Replacing the geofabric with a fme mesh geonet should 
alleviate the problem. 


45 



Section 4 

Data Quality Review 


r> 

This section presents the summarized results of QA 
procedures established to ensure the validity of the zinc 
and acute toxicity data collected during the demonstration. 
Section 4.1 discusses zinc data quality, and Section 4.2 
discusses acute toxicity data quality. A comprehensive 
discussion for both zinc and acute toxicity, along with 
supporting summary tables, is presented in the Technical 
Evaluation Report. 

4.1 Zinc Data Quality Review 

This section discusses the results of the QA procedures 
established to ensure the validity of the zinc data collected 
during the demonstration. The QA procedures were 
established prior to the demonstration and were recorded 
in the quality assurance project plan (Q APP) as part of the 
demonstration plan. Both field and analytical QA 
procedures were specified to ensure sample integrity and 
the generation of data of known quality. 

4.1.1 Quality Assurance Results for Field 
Sampling Activities 

The procedures followed during field activities to maintain 
sample integrity and quality are discussed below. They 
include specifications for sample collection, labeling, 
containerization, preservation, holding times, and chain of 
custody. 

Sample Containerization, Preservation, and Holding 
Times 

This section describes sample labeling, shipment, chain- 
of-custody, and laboratory receipt procedures for zinc 
samples. Conformance with and documentation of these 
procedures provide a definitive record of sample integrity 
from origin to analysis. 

Each sample container was labeled with a unique sample 
identification number. The label identified the sampling 
location, date, time of collection, and analysis to be 


performed. All chain-of-custody forms included the 
project number, project name, sampler’s name, station 
number, date, time, sampling location, number of containers, 
and analytical parameters. Samples were hand-delivered 
to Quanterra Environmental Services in Arvada, Colorado. 
Chain-of-custody forms gathered during the demonstration 
were reviewed for content and completeness and appeared 
in good order. 

All samples analyzed for critical parameters arrived at the 
laboratory intact. Several of the coolers used for shipping 
the samples arrived with inside temperatures greater than 
4 degrees Celsius as specified in the QAPP. However, 
the results of associated QA samples suggest that the 
elevated temperature did not affect sample integrity. All 
samples were analyzed within their designated holding 
times (6 months); the majority were analyzed within 
1 month of sample collection. 

Equipment and Field Blanks 

Equipment blanks were collected during the demonstration 
to assess sample contamination resulting from sampling 
equipment. Throughout the demonstration, dedicated 
sampling equipment was used for sample collection to 
reduce sample cross contamination. As a result, few 
equipment blanks or field blanks were collected during the 
demonstration. The data quality objective (DQO) for 
equipment and field blanks was results below reporting 
limits for all analytes. 

Two equipment blanks (WEV090794EB and EBO12197) 
were collected with a polyethelene dipper by pouring 
deionized water into the dipper and decanting the water 
into an appropriate sample container. The equipment 
blank collected in September 1994, contained an estimated 
zinc concentration of 0.019 mg/L, which is below the 
0.020 mg/L reporting limit. The equipment blank collected 
in January 1997, contained 0.052 mg/L zinc, above the 
0.020 mg/L reporting limit. 


46 



Field blanks were used to assess whether zinc 
contamination was introduced during the handling, 
presentation, or transport of aqueous samples. The field 
blank was prepared by adding deionized water into an 
appropriate sample container in place of a real sample. 

One field blank was collected during the demonstration 
(FB060194). Zinc was found in this field blank at a 
concentration of0.034 mg/L, slightly above the reporting 
limit of0.020 mg/L. 

The level of contamination in the equipment and field 
blanks qualifies data near the reporting limit for accuracy. 
The source of the contamination is unknown; however, 
the commercial distilled water is suspected. All of the 
CWS performance data contained zinc concentrations 
at least one order of magnitude greater than the 
reporting limit and in most cases two or three orders of 
magnitude above the reporting limit. Consequently, the 
demonstration zinc data are considered acceptable for 
their intended use. 

Method Blanks 


The QA objective for the CWS demonstration data were 
established in the QAPP with specific performance goals 
for precision, accuracy, representativeness, completeness, 
and comparability. The following sections evaluate the 
demonstration data with respect to these performance 
goals. 


Precision and Accuracy 


Precision is a measure of the reproducibility of 
measurements under a given set of conditions. Accuracy 
is the degree of agreement between an analytical 
measurement and the true value. The overall precision for 
zinc concentrations was a function of both sampling and 
laboratory precision. Overall precision was evaluated 
using data from field duplicates, and laboratory precision 
was evaluated using data from laboratory duplicates. 
Relative percent difference (RPD) between duplicate 
samples was used to evaluate precision using the following 
formula: 


RPD = J(A-B)| 
0.5 (A + B) 


X 100 


Method blanks verify that laboratory extraction and sample 
cleanup and concentration procedures used do not introduce 
contaminants that compromise the analytical results. 
Method blanks were prepared and analyzed with each 
batch of laboratory analysis. The method blank DQO was 
for results to be below reporting limits for all analytes of 
interest. 

Five out of the 40 batches analyzed during this 
demonstration contained reportable quantities of zinc in 
the method blanks. Values ranged from 0.020 mg/L to 
0.046 mg/L. All samples corresponding to these five 
analytical batches were qualified for blank contamination 
(B). All of the sample results were greater than five times 
the associated blank contamination; thus, no zinc results 
were qualified as nondetected due to blank contamination 
(UB). 

4.1.2 Quality Assurance Results for 
Sample Analysis 

Analytical QA includes methods and procedures used to 
ensure data reliability. This process involves establishing 
data quality objectives for the project data and developing 
data quality indicators (quanitative or qualitative measures 
of precision, accuracy, completeness, representativeness, 
and comparability) that can be used to determine whether 
the data meet the project’s QA objectives. 


where: A = first duplicate concentration 

B = second duplicate concentration or 

Fifteen field duplicate samples were collected during 
this demonstration, yielding RPDs ranging from 0 to 
3.7 percent. Laboratory duplicate control sampling were 
analyzed for 51 rounds of sampling activities. All laboratory 
RPDs were within the established DQO of 20 percent 
with the exception of one, of 28 percent. Overall, the 
precision objectives for zinc analyses were achieved. 

The accuracy of a measurement is affected by errors 
introduced through the sampling process and in handling, 
sample matrix, sample preservation, and analytical 
techniques. A program of sample spiking at the laboratory 
and analysis of standard reference materials (SRMs) was 
also used to evaluate laboratory accuracy. 

Accuracy for zinc measurements was estimated as percent 
recovery (%R) of the true analyte level from SRMs and 
by evaluation of matrix spike (MS) recoveries. The 
following formula was used to calculate MS percent 
recovery: 

% R = (S-C)/T X 100 


47 




where: S = measured spike concentration 
C = sample concentration 
T = true or actual concentration of the spike or 

MS spiking recoveries were all within the DQO limits with 
one exception. One MS sample analyzed (collected on 
July 27,1994) yielded a recovery of 134 percent, slightly 
above the DQO. When the data were rechecked by the 
laboratory, the deviations were not found to bias the 
results sufficiently to affect data use. The laboratory 
concluded that the magnitude of the errors was too small 
relative to the zinc concentrations to have a significant 
effect on the zinc values. 

Reported results for the SRM indicate that the analytical 
method measured larger concentrations of zinc than 
reported in National Institutes of Standards and Testing 
(NIST) standard reference material 1643c. The higher 
recoveries were considered to be the result of matrix 
interferences and the low level of zinc in the SRM. The 
DQO for accuracy is 75 to 125 percent recovery. SRM 
recoveries were 123 and 149 percent. Quanterra was 
immediately notified of the problem, and the laboratory 
control samples were checked to confirm that all other 
analytical controls were within acceptable parameters. 
Tetra Tech determined that some demonstration results 
with very low levels of zinc may be positively biased. The 
zinc results affected are from the upflow cell effluent 
during the first 6 months of operation. 

Overall laboratory accuracy for the demonstration data 
was acceptable. 

Representativeness 

Representativeness expresses the degree to which sample 
data accurately and precisely represent the characteristics 
of a population, parameter variations at a sampling point, 
or an environmental condition they are intended to 
represent. For the CWS demonstration, the low RPDs 
associated with field duplicate results suggest the data 
collected are representative of the CWS system for the 
environmental and physical conditions at the Burleigh 
Tunnel site. 

Completeness 

Completeness is a measure of the amount of acceptable 
data obtained compared to the amount of data needed 
to achieve a particular level of confidence in the results. 
Acceptable data are obtained when (1) samples are 
collected and analyzed in accordance with the 


QC procedures outlined in the demonstration plan, and 
(2) criteria that affect data quality are not exceeded. 
CWS percent project completeness (%C) was calculated 
using the following equation: 

%C = (V/T) X 100 

where: %C = percent completeness 

V = number of measurements judged 

acceptable 

T = total number of measurements planned 

The QA objective for degree of completeness was 
90 percent for the critical parameter zinc. All data 
collected are considered usable for the intended purpose; 
therefore, the QA objective for completeness was 
achieved. 

Comparability 

The comparability parameter is designed to identify 
deviations in the data that may result from inconsistencies 
in field conditions, sampling methods, or laboratory analysis. 
During this demonstration, changes in sampling techniques 
and laboratory analysis were minimized to ensure 
comparability of results. However, the end of the first 
SITE contract and delays in restarting the new SITE 
contract required the use of data collected by CDPHE. 
The results of a laboratory intercalibration exercise with 
Quanterra, the CDPHE laboratory (Analytica), and a 
referee laboratory suggest that the data are comparable. 

4.2 Acute Toxicity Data Quality Review 

This section discusses the results of QA data collected to 
document the validity of the acute toxicity data. The QA 
procedures were established prior to the demonstration 
and recorded in the QAPP as part of the demonstration 
plan. Both field and analytical QA procedures were 
specified to ensure sample integrity and the generation of 
data of known quality. 

4.2.1 Analytical Quality Assurance 

Analytical QA is the process of ensuring and confirming 
data reliability. This process includes establishing 
DQOs for the project data and developing data quality 
indicators (quantitative or qualitative measures of precision, 
accuracy, completeness, representativeness, and 
comparability) that can be used to evaluate whether the 
data met the project’s QA objectives. The QA objectives 
for acute toxicity testing during the CWS demonstration 


48 



were established in the QAPP and are summarized in the 
following discussions. 

Water Chemistry Results for Environmental 
Samples and Reference Toxicant Tests 

To ensure that laboratory water quality conditions did not 
adversely affect the reference toxicant or environmental 
sample results, water quality parameters were documented 
throughout all test series. The water chemistry results 
indicate that the water quality conditions for testing were 
appropriate for the test organisms during all test dates and 
that no abnormal water conditions were documented that 
could influence the survivability results. 

Precision and Accuracy 

Precision and accuracy in toxicity tests are controlled and 
evaluated through documentation of reference toxicant 
responses of indicator species against inter- and intra¬ 
laboratory historical records; and by carefully controlling 
and documenting the environmental conditions tested. 
The following discussion documents the laboratory testing 
conditions for growth, feeding, and maintenance of indicator 
species during the tests; and documents the results of 
indicator species survivability results against laboratory 
historical records for identical tests. 

Acute toxicity and metal concentration in the mine drainage 
were used to infer a response relationship between the 
most prevalent toxic component present (zinc) and indicator 
species survival. Preliminary chemical analysis had 
identified zinc in various forms as the most predominant 
metal contaminant. 

Zinc sulfate was used as a reference toxicant to simulate 
the population response of the indicator species to a 
soluble zinc compound present in the mine drainage 
matrix. Potassium chloride was used as a laboratory 
reference test for population viability and toxic response 
of the indicator species. 

Pimephales promelus and Ceriodaphnia dubia were used 
as the test organism populations in the 48-hour static- 
renewal acute toxicity tests. Indicator species survival 
rates (LC50) at the 95 percent confidence level (EPA 
1993a) in a static series of potassium chloride and zinc 
sulfate concentration dilutions were calculated and 
compared with laboratory historical records. The 
comparison provided a control on the viability of the test 
species and the testing methodology. 


The quantitative precision and accuracy requirements for 
acute toxicity for Pimephales promelus and Ceriodaphnia 
dubia when exposed to zinc sulfate were established by 
toxicant equivalent concentration values generated from 
both external and internal laboratory records of earlier 
tests. The quantitative precision and accuracy objectives 
for acute toxicity for Pimephales promelus and 
Ceriodaphnia dubia when exposed to potassium chloride 
were established by monthly cumulative laboratory toxicant 
equivalent concentration values. 

All reference toxicant results fell within the prescribed 
ranges, indicating that the response of the indicator 
species response to test conditions was appropriate for 
evaluating the toxin present. Therefore, the quantitative 
results of acute toxicity to the environmental samples are 
comparable to other tests under identical conditions. 

Sample Duplicates 

The results of sample (field) duplicates is another indicator 
of overall precision. The sample duplicate was collected 
on February 27, 1995 from the treated effluent from the 
downflow cell (samples designated WED and WED II). 

Generally, the analysis of duplicate acute toxicity values 
for sampling and analytical precision is a numerical 
comparison of the difference in reported acute toxicity 
values to the magnitude of the values themselves. 
However, sample WED for February 27, 1995 was not 
toxic enough to generate an LC50 value, which is the 
normal endpoint for acute toxicity analysis. Consequently, 
the analysis of test sampling and analytical precision 
presented is a subjective comparison of the sample and 
duplicate routine chemistry and intermediate toxicity 
results. 

The chemistry for duplicate samples WED and WEDII 
shows no significant difference, with less than 10 percent 
variation in all measured parameters. Those variables 
having the greatest difference - in pH, DO, and temperature 
- were consistently lower for WEDII than for WED. The 
values, however, do not strongly indicate a difference in 
water quality conditions. The initial and final chemistry for 
both species tests also show slight differences, but no 
consistent variability in an individual parameter. 

Qualitatively, the survival rates for C. dubia of the individual 
sample dilutions for duplicate samples WED and WEDII 
both show very slight toxicity, especially noting that both 
controls had survival rates of 20/20. Quantitatively, the 
100 percent WEDII sample yields a survival ratio 


49 




statistically different than the control when tested with 
Steel’s Many-One-Rank test at an = 0.05 (EPA 1993a). 
WED at 100 percent concentration did not exhibit sufficient 
mortality for the survival ratio to be statistically different 
than the control. 

The acute tests with P. promelas do not show any 
statistical difference from the control for WED or for 
WEDII; therefore, no toxicity for this species is evident. 
In general, C. dubia is more sensitive to environmental 
toxicants, so the absence of toxicity for P. promelas 
supports the presumption that WEDII is slightly toxic. 
Using the C. dubia results alone, it appears that there is a 
slight difference in the acute toxicity of the duplicate 
samples (WED and WEDII). Also, the arrival, initial, and 
final chemistry data show a difference in the characteristics 
in the ambient water between the two samples. Therefore, 
the duplicate analysis indicates that there is sufficient 
variability in the effluent stream to reflect a difference in 
the toxicity results of duplicate samples. However, this 
difference between duplicates is sufficiently small that the 
results of the acute toxicity tests, with LC50 as the 
endpoint, are not sensitive enough to calculate a coefficient 
of variation for effluent mine drainage samples. 

Representativeness 

For this project, representativeness for acute toxicity tests 
involved sample size, sampling times relative to seasonal 
temperature variation, and sampling locations. Most 
importantly, the changes due to seasonal environmental 
conditions needed to be documented to enable evaluation 
of zinc concentration reduction by biological conversion 
and uptake during cold stress conditions against warm 
temperature conditions. The QA goal was to obtain 
samples that represented biological water quality, measured 
by acute toxicity, in the treated and untreated mine 
drainage under typical seasonal environmental conditions. 
The primary seasonal environmental parameter of concern 
was temperature due to the regional extremes present at 
the demonstration location. 

Prior to the demonstration, it was known that three or four 
seasonal cycles would be required to conduct a statistical 
analysis of seasonal variation. The project budget and 
time schedule did not permit this type of data collection; 
consequently, the QA goal for representativeness 
was limited to successfully collecting data that would 
enable a limited evaluation of seasonal rise and fall 
of acute toxicity values in response to seasonal temperature 
stress. Since acute toxicity and zinc concentration data 
were obtained under environmental conditions 


representative of seasonal fluctuations in temperature in 
mine drainage influent and effluent, the QA objective for 
representativeness was met. 

Completeness 

Completeness is an assessment of the amount of valid 
data obtained from a measurement system compared to 
the amount of data expected to achieve a predefined 
quantity of information or level of confidence. The 
percent completeness is calculated by dividing the number 
of samples with acceptable data by the total number of 
samples planned to be collected and multiplying the result 
by 100. Greater than 90 percent completeness was 
achieved for all demonstration samples, and 100 percent 
of the critical samples for acute toxicity achieved 
acceptable results. 

Comparability 

The acute toxicity tests were conducted in accordance 
with the EPA guidance document “Methods for Measuring 
the Acute Toxicity of Effluents and Receiving Waters to 
Freshwater and Marine Organisms” (EPA 1991). All 
quality assurance guidance procedures have been adhered 
to, and the quantitative results for all QA criteria for 
reference toxicity fall within the specified limits. Therefore, 
the demonstration data are considered comparable to 
other acute toxicity data generated using these standard 
methods and adhering to the QA guidelines. 

4.3 Noncritical Parameters Data Quality 
Review 

Data quality review for the first noncrtical objective of 
substrate utilization, and the third noncritical objective of 
effluent impact to Clear Creek were included in the 
review for the number one critical obj ecti ve data. Analytical 
results for these two noncritical parameters were within 
the quality assurance objectives stated in the Demonstration 
Plan (PRC 1995). 

Data quality results for noncritical obj ecti ve number two, 
the metal removal by sulfate-reducing bacteria were 
within the parameters cited in the Demonstration Plan. 
As stated in the plan, the evaluation of sulfate-reduction 
was expected to be more qualitative in nature. Results for 
the bacteria counts and acid-volatile sulfides are considered 
acceptable quality. 

Specific data quality assurance objectives for the fourth, 
and final noncritical ojbective, compiled capital and 


50 



operating costs, were not stated in the Demonstration 
Plan. However, cost tracking and compilation was 
performed using a best professional judgment approach. 
These data are considered accurate and usable within 
accepted professional standards. 


51 




Section 5 

Economic Analysis 


This section presents cost estimates for using an anaerobic 
compost CWS system to treat mine drainage with water 
chemistry similar to the Burleigh Tunnel. The baseline 
scenario used for developing this cost estimate was a 50 
gpm flowrate, the total flow from the Burleigh Tunnel, and 
a 15-year system life. The baseline costs were then 
adjusted for flowrates of 25 gpm and 100 gpm to develop 
cost estimates for other cases. 

Cost estimates presented in this section are based primarily 
on data compiled during the SITE demonstration at the 
Burleigh Tunnel (CDPHE 1995). Additional cost data 
were obtained from standard engineering cost reference 
manuals (Means 1992). Costs have been assigned to 
11 categories applicable to typical cleanup activities at 
Superfund and RCRA sites (Evans 1990). Costs are 
presented in year 1995 dollars and are considered estimates, 
with an accuracy of plus 50 percent and minus 30 percent. 

5.1 Basis of Economic Analysis 

A number of factors affect the costs of treating mine 
drainage with an anaerobic compost CWS system. These 
factors generally include flow rate, type and concentration 
of contaminants, physical site conditions, geographical 
site location, and treatment goals. The characteristics of 
spent substrate produced by a CWS system will also 
affect disposal costs. Spent substrate will require off-site 
disposal. Mine drainage containing cadmium at 0.05 parts 
per million (ppm), iron at 50 ppm, nickel at 0.5 ppm, 
and zincat50 ppm was selected for this economic analysis. 
The following presents additional assumptions and 
conditions as they apply to each case. 

For each case, this analysis assumes that an up flow CWS 
system will treat contaminated mine drainage continuously, 
24 hours per day, 7 days per week. An average metals 
removal efficiency of 96 percent was assumed for all 
cases. Based on these assumptions, the CWS system will 


treat about 26.3 million gallons of water per year of 
operation at the baseline flowrate of 50 gpm. 

• Further assumptions about constructed wetlands 
treatment for each case include the following: 

• A residence time of 75 to 150 hours is recommended 
for adequate metals removal. 

• A porosity of 50 percent is assumed for the substrate 
material. 

• Two baseline wetlands, size of 90 feet by 90 feet by 
4 feet (2,300 cubic yards [yd 3 ]), will provide a 78 
hour residence time at a flowrate of 50 gpm (wetland 
size is directly proportional to flowrate). Square 
wetlands were used for the cost estimation; however, 
other shapes may be preferable. 

• Substrate material will require removal and 
replacement once every 5 years. 

• The spent substrate is not a RCRA hazardous waste: 
thus, it will be dewatered on site and can be recycled 
or disposed of at an industrial landfill. 

• An aerobic polishing pond to increase displaced 
oxygen is not required. 

This analysis assumes that aquatic-based standards are 
most appropriate; and the attainment of these standards 
depends on the affected organisms, receiving waters and 
volume of mine drainage. Attainment may not be feasible 
in all cases for the technology as tested during this 
demonstration. 

The following assumptions were also made for each case 
in this analysis: 

• The site is located within 200 miles of the disposal 
location. 


52 


The site is located within 100 miles of a moderate¬ 
sized city. 




• The site will allow for gravity flow of the mine 
drainage through the wetland. 

• A staging area is available for dewatering spent 
substrate. 

• Access roads exist at the site. 

• Utilities, such as electricity and telephone lines, are 
available on site. 

• The treatment goal for the site will be to reduce zinc 
contaminant levels by 90 percent. 

• Spent substrate will be dewatered and disposed of 
off site. 

• One influent water sample and two effluent water 
samples will be collected monthly and two composite 
substrate samples will be collected quarterly to 
monitor system performance. 

• One part-time operator will be required to inspect 
the system, collect all required samples, and conduct 
minor maintenance and repairs. 

5.2 Cost Categories 

Cost data associated with the CWS technology have been 
assigned to one of the following 11 categories: (1) site 
preparation; (2) permitting and regulatory requirements; 
(3) capital equipment and construction; (4) startup; 
(5) labor; (6) consumables and supplies; (7) utilities; 
(8) residual and waste shipping and handling; (9) analytical 
services; (10) maintenance and modifications; and 
(11) demobilization. Costs associated with each category 
are presented in the sections that follow. Some sections 
end with a summary of significant costs within the category. 
Table 9 presents the cost breakdown for the flow variant 
cases. This table also presents total one-time, fixed costs, 
and total variable O&M costs; the total project costs; and 
the costs per gallon of water treated. 

5.2.1 Site Preparation Costs 

Site preparation includes administration, pilot-scale testing, 
mobilization costs. This analysis assumes a total area of 
about 65 acres will be needed to accommodate the 
wetland and staging area, construction equipment, and 
sampling and maintenance equipment storage areas. A 
solid gravel (or ground) surface is preferred for any 
remote treatment project. Pavement is not necessary, but 
the surface must be able to support construction equipment. 
This analysis assumes adequate surface areas exist at the 
site and that only moderate modifications will be required 
for wetland construction. 


Administrative costs, such as legal searches and access 
rights, are estimated to be an additional $10,000. 

Mobilization involves transporting all construction 
equipment and materials to the site. For this analysis, it is 
assumed that the site is located within 100 miles of a city 
where construction equipment is available. The total 
estimated mobilization cost will be $5,000. 

For each case, total site preparation costs are estimated 
to be $15,000. 

5.2.2 Permitting and Regulatory 
Requirements 

Permitting and regulatory costs vary depending on whether 
treatment occurs at a Superfund site and on the disposal 
method selected for treated effluent and any solid wastes 
generated. At Superfund sites, remedial actions must be 
consistent with ARARs, environmental laws, ordinances, 
and regulations, including federal, state, and local standards 
and criteria. In general, ARARs must be identified on a 
site-specific basis. At an active mining site, a NPDES 
permit will likely be required and may require additional 
monitoring records and sampling protocols, which can 
increase permitting and regulatory costs. For this analysis, 
total permitting and regulatory costs are estimated to be 
$5,000. 

5.2.3 Capital Equipment 

Capital costs include all wetland construction and 
construction materials and a site building for housing 
sampling, monitoring, and maintenance equipment. 
Construction materials include sand, synthetic liners, 
geotextile liners, PVC piping, valves, concrete vaults or 
sumps, weirs, and other miscellaneous materials. Capital 
costs for the baseline wetland of 50 gpm are presented 
below. Site preparation and excavation include clearing 
the site of brush and trees, excavation of the wetland cell, 
grading the cell, and construction of the earthen berms. 
The total cost of site preparation and excavation is $ 19,500 
for the 50 gpm system. 

Construction of the wetland cell itself involves system 
design, subgrade preparation and installation of a sand 
layer, liner, piping distribution and collection systems, and 
the substrate. Also included is piping to and from the cell 
as well as system bypass piping, and concrete sumps with 
weirs at the influent of the wetland to control flow through 


53 



Table 9. CWS Costs for Different Treatment Flow Rates* 


System Life 15 Years 

Cost Categories 

25 gpm_50 gpm_100 gpm 


Fixed Costs 




Site Preparation 

Administrative 

Mobilization 

$15,000 

$10,000 

5,000 

$15,000 

$10,000 

5,000 

$15,000 

$10,000 

5,000 

Permitting and Regulatory 
Requirements 

$5,000 

$5,000 

$5,000 

Capital Equipment 

System Design 

Excavation and Site 
Preparation 

Wetland Cell Construction 

Piping and Valves 

Storage Building 

$215,300 

$50,000 

9,800 

120,000 

25,500 

10,000 

$345,000 

$50,000 

19.500 

240,000 

25.500 

10,000 

$604,500 

$50,000 

39,000 

480,000 

25,500 

10,000 

Startup 

$1,500 

$1,500 

$1,500 

Demobilization 

Excavation and Backfilling 
Substrate Disposal 

$52,250 

$10,000 

42,250 

$104,500 

$20,000 

84,500 

$209,000 

$40,000 

169,000 

Total Fixed Costs 

$316,000 

$492,000 

$844,000 

Variable Costs 




Labor 

Operations Staff 

$153,000 

$153,000 

$153,000 

$153,000 

$153,000 

$153,000 

Consumables and Supplies 

Personal Protective 
Equipment 

$39,000 

$39,000 

$39,000 

$39,000 

$39,000 

$39,000 

Utilities 

NA 

NA 

NA 

Residual and Waste Shipping and 
Handling 

Substrate Disposal 

$120,000 

40,000 (3) 

$240,000 

80,000 (3) 

$480,000 

160,000 (3) 

Analytical Services 

$360,000 

$360,000 

$360,000 

Maintenance and Modifications 

Annual Maintenance 

Substrate Removal and 
Replacement 

$247,550 

$5,000 

80,850 (3) 

$490,100 

$5,000 

161,700 (3) 

$975,200 

$5,000 

323,400 (3) 

Total Variable Costs 

$919,550 

$1,282,100 

$2,007,200 

Total Costs 

$1,235,500 

$1,774,100 

$2,851,200 

Total Cost Per Gallon Treated 

$0.0063 

$0.0045 

$0.0036 


*Costs are based on July 1995 dollars, rounded to the nearest $100. 
Substrate removal and replacement estimated to be necessary every 5 years. 
(3) N umber of removals anticipated 
NA Not applicable 


54 





























the system. The total cost for wetland cell construction of 
a 50 gpm system is $335,000. 

A small building is required for storing sampling equipment 
and providing work space for the system operator. The 
cost for a simple building with electricity has been estimated 
at $10,000. 

The total capital cost for a 50 gpm wetland system is 
$345,000. 

5.2.4 Startup 

Startup requirements are minimal for a wetland system. 
System startup involves introducing flow to the wetland 
with frequent inspections to verify proper hydraulic 
operation. Operators are assumed to be trained in health 
and safety procedures. Therefore, training costs are not 
incurred as a direct startup cost. The only costs directly 
related to system startup are labor costs associated with 
more frequent system inspection. Startup costs are 
estimated at $1,500. 

5.2.5 Labor 

Labor costs include a part-time technician to sample, 
operate, and maintain the system. Once the system is 
functioning, it is assumed to operate continuously at the 
design flow rate. One technician will monitor the system 
on a weekly basis. Weekly monitoring will require several 
hours 2 to 3 times per week to check flowrate and overall 
system operation. Sampling is assumed to be conducted 
once a month and will require two technicians for 2 hours. 
These requirements equate to 175 hours annually for 
general O&M. An additional 80 hours of labor are 
included for miscellaneous O&M and review of data. 
Based on $40 per hour for a technician, the annual cost for 
general labor O&M is $10,200. 

5.2.6 Consumables and Supplies 

The only consumables and supplies used during wetland 
operations are disposable PPE. Disposable PPE includes 
Tyvek coveralls, gloves, and bootcovers. The treatment 
system operator will wear PPE when required by health 
and safety plans during system operation. PPE will cost 
about $25 per day per person on site. Based on the 
assumed labor required above and an additional 22 days 
for miscellaneous O&M, PPE will be required 100 days 
annually, for an annual PPE cost of about $2,500. 


5.2.7 Utilities 

Utilities used by the wetland system are negligible. The 
wetland system requires no utilities for operation. The 
only utility required is for electricity for lights in the on-site 
storage building and for charging monitoring equipment. 
For this analysis, utility costs are assumed to be zero. 

5.2.8 Residual Waste Shipping and 
Handling 

The residual waste for the wetland is assumed to be spent 
substrate. This analysis assumes that substrate will 
require removal and replacement once every 5 years. It 
is assumed that spent substrate will be dewatered on site 
and disposed of at a recycler or landfill. Substrate removal 
and replacement costs are covered in Section 5.2.11, 
maintenance and modifications. Loading dewatered 
substrate into 20 yd 3 haul trucks is estimated to cost 
$14,500. Hauling the substrate to a recycler or landfill 
is estimated to cost $28,000; disposal of substrate at 
the landfill costs $42,000. Oversight of substrate removal, 
hauling and replacement is expected to cost $3,200 (10 8- 
hour days at $40/hr). Loading of the new substrate is 
expected to cost $12,000 and the cost of the substrate is 
$65,200. The total waste shipping and handling cost per 
substrate replacement is $161,700. Costs for residual 
waste shipping and handling are based solely on substrate 
volume. Costs for different sized wetlands are proportional 
to the 50 gpm baseline system described here. 

5.2.9 Analytical Services 

Analytical costs associated with a wetlands system include 
laboratory analysis, data reduction and tabulation, QA/ 
QC, and reporting. For each case, this analysis assumes 
that one influent sample and two effluent samples will be 
collected once a month and that two substrate samples 
will be collected quarterly. The substrate samples will be 
analyzed for total metals. Influent and effluent samples 
will be analyzed for total metals, ammonia, nitrate, 
phosphate, BOD, TSS, and TDS. Monthly laboratory 
analysis will cost about $1,050, and substrate analysis 
$3,500 per year. Data reduction, tabulation, Q A/QC, and 
reporting are estimated to cost about $660 per month. 
Total annual analytical services for each case are estimated 
to cost about $24,000 per year. 

5.2.10 Maintenance and Modifications 

Annual repair and maintenance costs are expected to be 
minimal and for this analysis are assumed to be $5,000 for 
each case. No modification costs are assumed to be 


55 



incurred. The major maintenance cost will be removal 
and replacement of the substrate every 5 years. Excavation 
of substrate material has been estimated to cost $14,500 
for the 50 gpm scenario. Replacement of the distribution 
and collection piping was estimated to cost $14,300. 
Purchase and transport of new substrate was estimated 
to cost $65,400. The total estimated cost of substrate 
removal and replacement is $161,700. The removal and 
replacement cost will vary proportionally with the wetland 
size. 

5.2.11 Demobilization 

Site demobilization costs include excavation of the substrate 
and concrete vaults and weirs, disposal of substrate, and 
backfilling the wetland. For the 50 gpm scenario, excavation 
costs are estimated at $10,000. Substrate disposal costs 
are $80,000. Backfilling of the wetland is expected to cost 
$ 10,000, assuming native material from the original wetland 
excavation was left on site. The total demobilization cost 
is estimated to be $104,500. This cost will vary 
proportionally with wetland size. 


56 



Section 6 
Technology Status 


Currently, several hundred constructed and natural 
wetlands are treating coal mine drainage in the eastern 
United States. The effectiveness of these systems is 
discussed in several publications including Hammer 1989, 
Moshiri 1993, and the proceedings of annual meetings of 
the American Society for Surface Mining and Reclamation, 
and several U.S. Bureau of Mines papers (U.S. Bureau 
of Mines Special Publication SP066-4 and Information 
Circular IC 9389) (see Appendix B). 

In addition, any constructed wetlands designed to treat 
metal mine drainages have been constructed and tested or 
are being tested by EPA, various state agencies, and 
industry. In Colorado, the state Division of Minerals has 
constructed several wetland systems to treat metal mine 
drainage. Constructed wetlands treatment is also being 
considered for the full-scale remedy of the Burleigh 
Tunnel drainage. 


57 



Section 7 
References 


Camp, Dresser, and McKee (CDM). 1993. Clear Creek 
Remedial Design Passive Treatment at Burleigh 
Tunnel, Draft Preliminary Design at Burleigh Tunnel. 
June. 

Colorado Department of Public Health and Environment 
(CDPHE). 1995. Facsimile Communication with 
Garry Farmer, Tetra Tech. February, 1995. 

Correns, C.W. 1969. Introduction to Mineralogy. 
Springer-Verlag. New York. Berlin. 

Environmental Restoration Unit Cost Book (ECHOS). 
1995. ECHOS, Los Angeles, California. 

Evans, G. 1990. Estimating Innovative Technology Costs 
for the SITE program. Journal of Air and Waste 
Management Association. 40:7:1047-1051. 

Gusek, J.J., and Wildeman, Dr. T. R.. 1995. New 

Developments in Passive Treatment of Acid Rock 
Drainage. Paper presented at Engineering Foundation 
Conference on Technological Solutions for Pollution 
Prevention in the Mining and Mineral Processing 
Industries, Palm Coast Florida, January 23, 1995. 

Gusek, J.J., J.T. Gormley, and J.W. Sheetz. 1994. 
Design and construction aspects of pilot-scale passive 
treatment systems for acid rock drainage at metal 
mines. Proc. Society of Chemical Industry 
Symposium. Chapman and Hall, London. 

Hammer, D.A. 1989. Constructed Wetlands 
for Wastewater Treatment. Lewis Publishers. 
Chelsea, Michigan. 

Hedin, R.S., R.W. Narin, and R.L.P. Kleinmann. 1994. 
Passive Treatment of Coal Mine Drainage. United 
States Bureau of Mines Information Circular 9389. 

Klusman, R.W. 1993. Computer Code to Model 
Constructed Wetlands for Aid in Engineering Design. 
Report to United States Bureau of Mines, Contract 
J0219003. 

Means, R.S. 1992. Means Building Construction 
Cost Data. Construction Consultants and Publishers, 
Kingston, Massachusetts. 


Metcalf and Eddy, Inc. 1979. Wastewater Engineering 
Treatment, Disposal, and Reuse. Revised by George 
Tchobanoglous and Franklin L. Burton. McGraw- 
Hill Publishing Company. New York, New York. 

Moshiri, G.A. 1993. Constructed Wetlands for 
Water Quality Improvement. Lewis Publishers. 
Boca Raton, Florida. 

PRC Environmental Management, Inc. (PRC) 1993. 
Colorado Department of Public Health and 
Environment, Constructed Wetlands System 
Treatability Study at the Burleigh Tunnel, Silver 
Plume, Colorado, Treatability Study Work Plan, 
Denver, Colorado, February 1993. 

PRC. 1995. Colorado Department of Public Health and 
Environment Constructed Wetlands System 
Demonstration Plan. July. 

Reynolds, J.S. 1991. Determination of the Rate of 
Sulfide Production by Sulfate-reducing Bacteria at 
the Big 5 Wetland. Masters Thesis. Colorado 
School of Mines. Golden, Colorado. 

U.S. Bureau of Mines. 1994b. Proceedings of the 
International Land Reclamation and Mine Drainage 
Conference and Third International Conference on 
the Abatement of Acidic Drainage. Pittsburgh, 
Pennsylvania, April 24-29, 1994, Bureau of Mines 
Special Publication SP 066-4. 

U.S. Environmental Protection Agency (EPA). 1988. 
Constructed Wetlands and Aquatic Plant Systems 
for Municipal Wastewater Office of Research and 
Development. Washington, D.C. EPA/625/1-88/ 
022. September. 

EPA. 1993a. Methods for Measuring the 
Acute Toxicology of Effluents and Receiving Waters 
to Freshwater and Marine Organisms. Office of 
Research and Development. Washington, D.C. 
EPA/600/4-90/027F. 4th Edition. September. 

EPA. 1993b. Handbook for Constructed 
Wetlands Receiving Acid Mine Drainage. Office of 
Research and Development. Cincinnati, OH. 
September. 


58 



Appendix A 

Analytical Results Summary Tables 


59 






Table A-l. Influent Results 


INFLUENT 





WI030994 

WI032394 

WI040694 

WI042094 

WI050594 

W1051994 



ANALYTICAL 

03/09/94 

03/23/94 

04/06/94 

04/20/94 

05/05/94 

05/19/94 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

ND 

ND 

ND 

ND 

0.045 


ARSENIC 

6020 

ND 

0.0041 

0.0068 

0.020 

0.060 

0.052 


CADMIUM 

6020 

0.10 

0.099 

0.10 

0.10 

0.098 

0.081 


CALCIUM 

6010 

84.8 

88.0 

91.7 

96.9 

89.9 

83.2 


IRON 

6010 

0.31 

0.33 

0.33 

0.34 

0.32 

0.21 


LEAD 

6020 

0.014 

0.015 

0.014 

0.016 

0.016 

0.014 


MAGNESIUM 

6010 

41.8 

43.1 

44.2 

46.5 

47.1 

49.1 


MANGANESE 

6010 

2.3 

2.4 

2.5 

2.6 

2.3 

1.8 


NICKEL 

6010 

0.045 

0.039 

0.042 

0.047 

0.043 

0.035 


POTASSIUM 

6010 

2.6 

2.9 

3.0 

3.1 

3.6 

3.2 


SILVER 

6020 

0.0011 

0.00012 

0.000066 

0.000070 

0.000098 

0.00019 


SODIUM 

6010 

10.3 

9.3 

10.9 

9.1 

14.0 

10.5 


ZINC 

6010 

55.0 

56.1 

60.1 

64.0 

56.1 

44.8 


ANIONS: 









SULFATE 

300.0 

386 

374 

387 

384 

317 

314 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

1.0 

1.2 

1.1 

1.1 

0.98 

1.0 


CHLORIDE 

300.0 

19.9 

21.8 

22.3 

21.9 

19.0 

15.0 


PHOSPHORUS, TOTAL 

365.3 

ND 

ND 

ND 

ND 

ND 

ND 


ORTHOPHOSPHATE 

365.3 

ND 

0.30 

ND 

ND 

ND 

0.40 


NITRATE PLUS NITRITE ASN 

353.2 

ND 

ND 

0.060 

0.11 

ND 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

0.060 

0.11 

ND 

ND 


AMMONIA 

350.1 

ND 

ND 

ND 

ND 

ND 

ND 


TOTAL SOLIDS: 









TSS 

160.2 

16.8 

8.8 

20.4 

15.2 

7.4 

8.4 


TDS 

160.1 

732 

655 

640 

663 

641 

622 


TOC 

9060 

1.1 

NA 

NA 

ND 

NA 

NA 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

100 

107 

105 

107 

104 

107 


ALKALINITY, BICARB 









AS CAC03 

310.1 

100 

107 

105 

107 

104 

107 


DISSOLVED OXYGEN (mg/L) 

— 

8.1 

8.3 

* 

oo 

vb 


NA 

NA 


PH 

— 

7.4 

7.5 

7.5 


7.4 

7.5 


CONDUCTIVITY (pS) 

-- 

730 

745 

745 


699 

698 


TEMPERATURE (degrees C) 

- 

6.9 

7.3 

7.3 


8.9 

9.4 


— = Not applicable NA = Not analyzed 

pS = MicroSiemens ND = Not detected 

mgL = Milligrams per liter 


60 




















Table A-l (continued). Influent Results 


INFLUENT 





WI060194 

WI062994 

WI071394 

WJ072894 

WI081594 

WI082494 



ANALYTICAL 

06/01/94 

06/29/94 

07/13/94 

07/28/94 

08/15/94 

08/24/94 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

0.068 

ND 

ND 

ND 

ND 


ARSENIC 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CADMIUM 

6020 

0.092 

0.089 

0.086 

0.098 

0.10 

0.0952 


CALCIUM 

6010 

89.6 

86.1 

94.5 

91.2 

92.5 

94.6 


IRON 

6010 

0.25 

0.23 

0.23 

0.30 

0.24 

0.25 


LEAD 

6020 

0.020 

0.017 

0.013 

0.017 

0.016 

0.014 


MAGNESIUM 

6010 

50.6 

45.4 

48.3 

46.4 

47.7 

48.1 


MANGANESE 

6010 

1.9 

2.1 

2.2 

2.2 

2.3 

2.4 


NICKEL 

6010 

0.033 

0.045 

0.044 

0.043 

0.042 

0.046 


POTASSIUM 

6010 

3.6 

3.0 

3.1 

2.9 

2.9 

3.2 


SILVER 

6020 

0.00019 

ND 

0.00013 

0.00015 

0.00017 

ND 


SODIUM 

6010 

13.2 

12.8 

13.00 

12.0 

14.4 

15.3 


ZINC 

6010 

49.1 

54.2 

56.8 

59.1 

54.7 

57.5 


ANIONS: 









SULFATE 

300.0 

357 

378 

377 

397 

374 

403 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

1.0 

1.0 

0.90 

1.1 

1.1 

1.1 


CHLORIDE 

300.0 

16.9 

17.9 

17.5 

18.7 

18.6 

19.6 


PHOSPHORUS, TOTAL 

365.3 

ND 

ND 

ND 

ND 

ND 

ND 


ORTHOPHOSPHATE 

365.3 

ND 

0.44 

ND 

0.077 

ND 

ND 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

ND 

2.0 

1.7 

1.9 


NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

2.0 

1.7 

1.9 


AMMONIA 

350.1 

ND 

ND 

ND 

ND 

ND 

ND 


TOTAL SOLIDS: 









TSS 

160.2 

4.4 

11.2 

9.2 

9.6 

2.4 

18.4 


TDS 

160.1 

657 

680 

685 

707 

759 

703 


TOC 

9060 

NA 

NA 

NA 

NA 

NA 

NA 


ALKALINITY, TOTAL: 

AS CaC03 

ALKALINITY, BICARB 

310.1 

109 

107 

109 

103 

105 

102 


ASCAC03 

310.1 

109 

107 

109 

103 

105 

102 


DISSOLVED OXYGEN (mg/L) 

- 

8.7 

NA 

8.2 

NA 

NA 

7.6 


pH 

— 

7.6 

7.57 

7.5 

NA 

7.5 

7.4 


CONDUCTIVITY (pS) 

-- 

775 

980 

950 

927 

948 

920 


TEMPERATURE (degrees C) 

- 

9.4 

9.5 

9.4 

9.5 

9.4 

9.4 


** = Degrees Farenheit NA = Not analyzed 

— = Not applicable ND = Not detected 

pS = micro Siemens 
mg/L= Milligrams per liter 


61 



















Table A-l (continued). Influent Results 




I 

NFLUEN1 







ANALYTE 

ANALYTICAL 

METHOD 

WI090794 

09/07/94 

mg/L 

WI091994 

09/19/94 

mg/L 

WI100494 

10/04/94 

mg/L 

WI101994 

10/19/1994 

mg/L 

Wll 10294 

11/02/94 

mg/L 

W1112094 

11/20/94 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

ND 

ND 

ND 

0.030 

ND 


ARSENIC 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CADMIUM 

6020 

0.098 

0.085 

0.089 

.10 

0.10 

0.091 


CALCIUM 

6010 

90.2 

89.7 

92.6 

92.4 

89.2 

93.5 


IRON 

6010 

0.29 

0.29 

0.31 

0.25 

0.28 

0.32 


LEAD 

6020 

0.017 

0.015 

0.014 

0.014 

0.014 

0.016 


MAGNESIUM 

6010 

46.5 

46.6 

47.3 

46.7 

46.2 

47.3 


MANGANESE 

6010 

2.3 

2.3 

2.3 

2.4 

2.2 

2.3 


NICKEL 

6010 

0.047 

0.042 

0.052 

0.046 

0.051 

0.050 


POTASSIUM 

6010 

3.9 

3.1 

3.0 

3.0 

2.9 

3.1 


SILVER 

6020 

0.00040* 

0.00041 

0.00050 

ND 

ND 

0.00030 


SODIUM 

6010 

12.1 

12.5 

11.6 

13 

14.8 

14.4 


ZINC 

6010 

56.4 

57.6 

59.7 

57.6 

56.5 

58.2 


ANIONS: 

SULFATE 

300.0 

416 

404 

400 

409 

410 

407 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

1.0 

1.0 

1.0 

ND 

1.0 

1.1 


CHLORIDE 

300.0 

20.2 

19.6 

19.8 

19.5 

20.1 

21.3 


PHOSPHORUS, TOTAL 

365.3 

ND 

ND 

ND 

ND 

ND 

ND 


ORTHOPHOSPHATE 

365.3 

ND 

ND 

ND 

ND 

0.13 

ND 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

ND 

ND 

ND 

ND 


NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

ND 

ND 

ND 

ND 

ND 

ND 


TOTAL SOLIDS: 

TSS 

160.2 

17.6 

8.4 

18.8 

18.8 

8.0 

18.0 


TDS 

160.1 

711 

723 

695 

695 

709 

711 


TOC 

9060 

NA 

NA 

NA 

NA 

NA 

NA 


ALKALINITY, TOTAL: 

AS CaC03 

310.1 

102 

101 

112 

102 

82.4 

101 


ALKALINITY, BICARB 

AS CAC03 

310.1 

102 

101 

112 

102 

82.4 

101 


DISSOLVED OXYGEN (mg/L) 

- 

9.5 

7.8 

NA 

NA 

NA 

NA 


pH 

- 

7.41 

7.4 

7.4 

7.1 

6.9 

6.9 


CONDUCTIVITY (pS) 

-- 

922 

930 

935 

750 

900 

NA 


TEMPERATURE (degrees C) 

-- 

9.3 

9.3 

9.1 

8.5 

8.7 

8.1 


— = Not applicable NA = Not detected 

pS = MicroSicmcns ND = Not detected 

mgT, = Milligrams per liter 


62 






















Table A-l (continued). Influent Results 


INFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WI113094 

11/30/94 

mg/L 

WI121494 

12/14/94 

mg/L 

WI010495 

01/04/95 

mg/L 

WI011895 

01/18/95 

mg/L 

WI020195 

02/01/95 

mg/L 

WI021595 

02/15/95 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

0.036 

0.032 

0.038 

0.047 

0.043 


ARSENIC 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CADMIUM 

6020 

0.086 

0.092 

0.82 

0.076 

0.089 

0.084 


CALCIUM 

6010 

95.4 

98.1 

87.7 

90.8 

90.1 

100.0 


IRON 

6010 

0.34 

0.37 

0.31 

ND 

0.34 

0.39 


LEAD 

6020 

0.014 

0.018 

0.016 

0.015 

0.016 

0.015 


MAGNESIUM 

6010 

47.7 

48.9 

46.5 

45.4 

44.1 

49.4 


MANGANESE 

6010 

2.5 

2.5 

2.3 

2.4 

2.4 

2.7 


NICKEL 

6010 

0.044 

0.050 

0.048 

0.046 

0.052 

0.048 


POTASSIUM 

6010 

2.8 

3.3 

2.9 

3.0 

2.8 

3.5 


SILVER 

6020 

0.00036 

ND 

0.00037 

0.00021 

ND 

ND 


SODIUM 

6010 

14.2 

19.5 

15.0 

15.9 

14.1 

20.4 


ZINC 

6010 

62.8 

63.0 

55.5 

57.1 

56.6 

58.9 


ANIONS: 









SULFATE 

300.0 

411 

413 

395 

386 

402 

390 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

1.1 

1.0 

1.1 

1.1 

1.1 

1.1 


CHLORIDE 

300.0 

21.4 

21.2 

21.6 

21.7 

22.5 

22.8 


PHOSPHORUS, TOTAL 

365.3 

ND 

ND 

ND 

ND 

ND 

ND 


ORTHOPHOSPHATE 

365.3 

0.13 

0.36 

ND 

ND 

ND 

0.10 


NITRATE PLUS NITRITE AS N 

353.2 

ND 

ND 

ND 

ND 

1.7 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

1.7 

ND 


AMMONIA 

350.1 

ND 

ND 

ND 

ND 

ND 

ND 


TOTAL SOLIDS: 









TSS 

160.2 

16.4 

10.4 

5.2 

12.0 

12.8 

12.8 


TDS 

160.1 

711 

687 

689 

693 

694 

656 


TOC 

9060 

NA 

NA 

NA 

NA 

NA 

NA 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

99.6 

103 

104 

106 

106 

106 


ALKALINITY, BICARB 









AS CAC03 

310.1 

99.6 

103 

104 

106 

106 

106 


DISSOLVED OXYGEN (mg/L) 

— 

NA 

8.0 

8.5 

7.3 

7.6 

NA 


pH 

-- 

6.9 

7.54 

7.5 

7.5 

7.9 

7.0 


CONDUCTIVITY (pS) 

-- 

605 

600 

610 

600 

610 

NA 


TEMPERATURE (degrees C) 

-- 

7.9 

8.0 

6.5 

9.0 

7.9 

8.1 


* = Dissolved metals NA = Not analyzed 

— = Not applicable ND = Not detected 

pS= Microsiemens 
mg/L = Milligrams per liter 


63 



















Table A-l (continued). Influent Results 


INFLUENT 





WI022795 

W1031595 

WI032995 

W1041295 

WI042695 

WI051095 



ANALYTICAL 

02/27/95 

03/15/95 

03/29/95 

04/12/95 

04/26/95 

05/10/95 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.024 

0.049 

ND 

ND 

0.060 

0.15 


ARSENIC 

f> 

CADMIUM 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


6020 

0.071 

0.076 

0.074 

0.057 

0.095 

0.095 


CALCIUM 

6010 

92.6 

91.4 

85.2 

90.9 

88.2 

92.0 


IRON 

6010 

0.33 

0.36 

0.33 

0.32 

0.41 

0.48 


LEAD 

6020 

0.014 

0.016 

0.014 

0.015 

0.022 

0.026 


MAGNESIUM 

6010 

45.1 

44.4 

41.9 

42.9 

41.2 

41.9 


MANGANESE 

6010 

2.5 

2.5 

2.3 

2.4 

2.6 

3.0 


NICKEL 

6010 

0.068 

0.045 

0.045 

0.048 

0.071 

0.054 


POTASSIUM 

6010 

2.9 

2.9 

2.8 

3.0 

2.9 

3.1 


SILVER 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


SODIUM 

6010 

16.2 

15.8 

16.4 

16.1 

14.2 

14.8 


ZINC 

6010 

58.6 

57.0 

53.1 

55.0 

55.7 

61.4 


ANIONS: 









SULFATE 

300.0 

384.0 

384.0 

368.0 

376.0 

370.0 

374 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

1.1 

1.1 

1.0 

1.0 

1.1 

1.1 


CHLORIDE 

300.0 

22.6 

22.4 

23.1 

22.4 

23.8 

20.5 


PHOSPHORUS, TOTAL 

365.3 

ND 

ND 

ND 

ND 

ND 

ND 


ORTHOPHOSPHATE 

365.3 

ND 

ND 

ND 

0.11 

ND 

ND 


NITRATE PLUS NITRITE ASN 

353.2 

ND 

ND 

ND 

ND 

0.14 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

0.14 

ND 


AMMONIA 

350.1 

ND 

ND 

ND 

ND 

ND 

ND 


TOTAL SOLIDS: 









TSS 

160.2 

11.2 

9.2 

12.8 

14.4 

7.2 

2.8 


TDS 

160.1 

692 

672 

655 

656 

575 

689 


TOC 

9060 

NA 

NA 

NA 

NA 

NA 

ND 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

107 

104 

107 

107 

104 

103 


ALKALINITY, BICARB 









AS CAC03 

310.1 

107 

104 

107 

107 

104 

103 


DISSOLVED OXYGEN (mg/L) 

— 

7.8 

NA 

7.5 

8.6 

7.5 



pH 

-- 

7.4 

7.5 

7.7 

7.5 

NA 



CONDUCTIVITY (pS) 

- 

630 

620 

600 

620 

600 



TEMPERATURE (degrees C) 

- 

8.6 

9.3 

8.1 

8.4 

9.0 



* = Dissolved metals NA = Not analyzed 

-- = Not applicable ND = Not detected 

pS = Microsiemens 
mg/L = Milligrams per liter 


64 




















Table A-l (continued). Influent Results 


INFLUENT 





VVI061295 

WI062895 

WI071095 

WI072695 

WI080895 

WI082395 



ANALYTICAL 

6/12/1995 

6/28/1995 

7/10/1995 

7/26/1995 

8/8/1995 

8/23/1995 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.065 

ND 

ND 

ND 

ND 

0.079 


ARSENIC 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CADMIUM 

6020 

0.25 

0.26 

0.25 

0.24 

0.26 

0.240 


CALCIUM 

6010 

94.4 

111 

119 

129 

123 

125 


IRON 

6010 

0.12 

0.11 

0.10 

ND 

0.15 

0.19 


LEAD 

6020 

0.058 

0.051 

0.050 

0.038 

0.043 

0.039 


MAGNESIUM 

6010 

58.3 

61.4 

64.0 

64.2 

61.7 

61.3 


MANGANESE 

6010 

3.9 

4.4 

5.0 

5.5 

5.2 

5.2 


NICKEL 

6010 

0.061 

0.073 

0.081 

0.084 

0.093 

0.086 


POTASSIUM 

6010 

4.1 

ND 

3.6 

3.7 

3.5 

3.2 


SILVER 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


SODIUM 

6010 

9.9 

14.2 

14.8 

13.2 

14.1 

15.2 


ZINC 

6010 

75.5 

86.8 

99.8 

105 

109 

108 


ANIONS; 









SULFATE 

300.0 

499 

502 

582 

596 

638 

630 


SULFIDE TOTAL 

376.2 








FLUORIDE 

340.2 

0.8 

0.89 

0.96 

0.88 

0.87 

0.95 


CHLORIDE 

300.0 

6.9 


8.8 

10.2 

11.7 

13.1 


PHOSPHORUS, TOTAL 

365.3 

ND 

ND 

ND 

ND 

ND 

0.093 


ORTHOPHOSPHATE 

365.3 

ND 

ND 

ND 

ND 

0.095 

ND 


NITRATE PLUS NITRITE ASN 

353.2 

0.13 

0.10 

ND 

0.63 

ND 

ND 


NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

0.13 

ND 

ND 

0.63 

ND 

ND 


AMMONIA 

350.1 

ND 

ND 

ND 

ND 

ND 

ND 


TOTAL SOLIDS; 









TSS 

160.2 

20.4 

20.4 

24.8 

22.4 

18.8 

32.0 


TDS 

160.1 

838 

967 

1010 

999 

10.0 

1050 


TOC 

9060 








ALKALINITY, TOTAL; 

AS CaC03 

310.1 

120 

125 

118 

107 

107 

107 


ALKALINITY, BICARB 

AS CAC03 

310.1 

120 

125 

118 

107 

107 

107 


DISSOLVED OXYGEN (mg/L) 

— 

NA 

7.1 

NA 

NA 

NA 

NA 


pH 

- 

7.4 

7.2 

7.4 

NA 

NA 

NA 


CONDUCTIVITY (pS) 

- 

NA 

700 

NA 

NA 

750 

NA 


TEMPERATURE (degrees Q 

- 

10.2 

10.3 

10.3 

NA 

10.4 

NA 


* = Dissolved metals NA = Not analyzed 

— = Not applicable ND = Not detected 

pS = Microsiemens 
mgL = Milligrams per liter 


65 





















Table A-l (continued). Influent Results 


INFLUENT 



ANALYTE 

ANALYTICAL 

WI090595 

WI110995 

CDPHE 

CDPHE 

CDPHE 

CDPHE 



METHOD 

9/5/1995 

11/9/1995 

1/29/1996 

2/29/1996 

4/25/1996 

5/31/1996 




mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

ND 

NA 

NA 

NA 

NA 


ARSENIC 

6020 

ND 

ND 

NA 

NA 

NA 

NA 


CADMIUM 

CALCIUM 

6020 

0.24 

0.20 

0.160 

0.200 

0.12 

0.14 


6010 

123 

113 

NA 

NA 

NA 

NA 


IRON 

6010 

0.28 

0.18 

0.24 

0.26 

0.18 

0.17 


LEAD 

6020 

0.038 

0.027 

NA 

NA 

NA 

NA 


MAGNESIUM 

6010 

60.2 

56.2 

NA 

NA 

NA 

NA 


MANGANESE 

6010 

5.2 

5.2 

3.60 

3.50 

2.4 

2.7 


NICKEL 

6010 

0.087 

0.082 

NA 

NA 

NA 

NA 


POTASSIUM 

6010 

ND 

3.2 

NA 

NA 

NA 

NA 


SILVER 

6020 

ND 

ND 

NA 

NA 

NA 

NA 


SODIUM 

6010 

12.4 

15.6 

NA 

NA 

NA 

NA 


ZINC 

6010 

107 

105 

73 

69 

46 

56 


ANIONS: 









SULFATE 

300.0 

652 

591 

490 

450 

NA 

NA 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

0.88 

0.97 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

NA 

17.7 

NA 

NA 

NA 

NA 


PHOSPHORUS, TOTAL 

365.3 

0.067 

0.060 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

ND 

0.20 

NA 

NA 

NA 

NA 


NITRATE PLUS NITRITE AS N 

353.2 

ND 

ND 

NA 

NA 

NA 

NA 


NITRITE AS N 

354.1 

ND 

ND 

NA 

NA 

NA 

NA 


NITRATE ASN 

353.2/354.1 

ND 

ND 

NA 

NA 

NA 

NA 


AMMONIA 

350.1 

ND 

ND 

NA 

NA 

NA 

NA 


TOTAL SOLIDS: 









TSS 

160.2 

18.4 

14.4 

NA 

NA 

NA 

NA 


TDS 

160.1 

1050 

956 

NA 

NA 

NA 

NA 


TOC 

9060 




NA 

NA 

NA 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

107 

95.7 

NA 

NA 

NA 

NA 


ALKALINITY, BICARB 




NA 

NA 

NA 

NA 


ASCAC03 

310.1 

107 

95.7 

NA 

NA 

NA 

NA 


DISSOLVED OXYGEN (mg/L) 

— 



NA 

NA 

NA 

NA 


pH 

- 



NA 

NA 

NA 

NA 


CONDUCTIVITY (pS) 

- 



NA 

NA 

NA 

NA 


TEMPERATURE (degrees C) 

- 



NA 

NA 

NA 

NA 


* = Dissolved metals NA = Not analyzed 

— = Not applicable ND = Not detected 

pS = Micro siemens 
mg/L = Milligrams per liter 


66 




















Table A-l (continued). Influent Results 


INFLUENT 





CDPHE 

CDPHE 

CDPHE 

WI120996 

WI012197 

WI022097 



ANALYTICAL 

6/14/1996 

7/19/1996 

8/31/1996 

12/9/1996 

1/21/1997 

2/20/1997 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

NA 

NA 

NA 

ND 

ND 

ND 


ARSENIC 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


CADMIUM 

6020 

0.16 

0.19 

0.20 

0.15 

0.12 

0.11 


CALCIUM 

6010 

NA 

NA 

NA 

104 

100.0 

105 


IRON 

6010 

0.18 

0.20 

0.24 

0.30 

0.30 

0.33 


LEAD 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


MAGNESIUM 

6010 

NA 

NA 

NA 

52.8 

51.2 

52 


MANGANESE 

6010 

2.9 

3.5 

4.1 

3.7 

3.5 

3.7 


NICKEL 

6010 

NA 

NA 

NA 

0.07 

0.06 

0.06 


POTASSIUM 

6010 

NA 

NA 

NA 

3.1 J 

3.0 J 

3.0 J 


SILVER 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


SODIUM 

6010 

NA 

NA 

NA 

17.4 

16.4 

17.0 


ZINC 

6010 

60 

71 

84 

78 

74 

78 


ANIONS: 









SULFATE 

300.0 

430 

490 

520 

488 

491 

471 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

NA 

NA 

NA 

17.8 

18.2 

18.3 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

NA 

NA 

NA 

0.31 

0.17 

0.22 


NITRATE PLUSNITRITE ASN 

353.2 

NA 

NA 

NA 

ND 

ND 

ND 


NITRITE ASN 

354.1 

NA 

NA 

NA 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

NA 

NA 

NA 

ND 

ND 

ND 


AMMONIA 

350.1 

NA 

NA 

NA 

ND 

ND 

ND 


TOTAL SOLIDS 









TSS 

160.2 

NA 

NA 

NA 

NA 

8.4 

3.2 


TDS 

160.1 

NA 

NA 

NA 

849 

796 

809 


TOC 

9060 

NA 

NA 

NA 

0.8J 

1.1 

1.8 


ALKALINITY, TOTAL: 









ASCaC03 

310.1 

NA 

NA 

NA 

97.6 

94.9 

101 


ALKALINITY, BICARB 


NA 

NA 

NA 





ASCAC03 

310.1 

NA 

NA 

NA 

97.6 

94.9 

101 


DISSOLVED OXYGEN (mg/L) 

— 

NA 

NA 

NA 

7.4 

8.8 

8.6 


pH 

— 

NA 

NA 

NA 

7.2 

5.1 

7.5 


CONDUCTIVITY (pS) 

-- 

NA 

NA 

NA 

NA 

NA 

NA 


TEMPERATURE (degrees Q 

— 

NA 

NA 

NA 

10.0 

8.2 

3.2 


* = Dissolved metals NA = Not analyzed 

— = Not applicable ND = Not detected 

pS = Microsiemens 
mg / L= Milligrams per liter 


67 



















Table A-l (continued). Influent Results 


INFLUENT 





WI032097 

WI042297 

WI052897 

WI062397 

WI082897 

WI093097 



ANALYTICAL 

3/20/1997 

4/22/1997 

5/28/1997 

6/23/1997 

8/28/1997 

9/30/1997 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

0.17 

ND 

ND 

ND 

ND 


ARSENIC 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


CADMIUM 

6020 

0.14 

0.07 

0.11 

0.19 

0.22 

0.200 


CALCIUM 

6010 

97.5 

67.2 

86.4 

95.6 

121 

119 


IRON 

6010 

0.34 

0.34 

0.24 

0.26 

0.3 

0.33 


LEAD 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


MAGNESIUM 

6010 

48.8 

37.3 

53.8 

52.3 

61.9 

58.4 


MANGANESE 

6010 

3.6 

2.0 

2.7 

3.3 

4.9 

4.9 


NICKEL 

6010 

0.07 

0.034 J 

0.042 

0.030 J 

0.090 

0.098 


POTASSIUM 

6010 

ND 

2.7 J 

3.3 J 

3.5 J 

4.8 J 

3.4 J 


SILVER 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


SODIUM 

6010 

15.6 

ND 

14.9 

ND 

ND 

18.3 


ZINC 

6010 

75 

42 

56 

72 

104 

104 


ANIONS: 









SULFATE 

300.0 

476 

279 

358 

428 

541 

568 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

18.7 

9.3 

7.2 

9.2 

13.8 

16 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

0.15 

ND 

ND 

0.10 

ND 

ND 


NITRATE PLUS NITRITE ASN 

353.2 

ND 

0.14 

ND 

0.14 

ND 

0.19 


NITRITE AS N 

354.1 

0.0021 J 

0.0046 J 

0.0024 J 

0.0028 J 

0.0037 J 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

ND 

ND 

ND 

ND 

ND 

ND 


TOTAL SOLIDS: 









TSS 

160.2 

7.6 

1.6 J 

12.4 

14.4 


16.4 


TDS 

160.1 

751 

507 

653 

765 

927 

940 


TOC 

9060 

0.20 J 

1.30 

1.4 

0.98 J 

0.80 J 

0.58 J 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

96.3 

99.7 

107 

121 


102 


ALKALINITY, BICARB 









AS CAC03 

310.1 

96.3 

99.7 

107 

121 


102 


DISSOLVED OXYGEN (mg/L) 

— 

7.8 

7.3 

7.3 

8 

8.7 

NA 


pH 

- 

6.9 

7.4 

7.4 

7.5 

6.9 

6.9 


CONDUCTIVITY (pS) 

-- 

NA 

NA 

NA 

NA 

NA 

NA 


TEMPERATURE (degrees C) 

- 

8.6 

9.7 

10.5 

9.7 

9.6 

9.4 


* = Dissolved metals NA = Not analyzed 

— = Not applicable ND = Not detected 

pS = Micro siemens 
mg/L = Milligrams per liter 


68 




















Table A-l (continued). Influent Results 


INFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WI102997 

10/29/1997 

mg/L 

WT112597 

11/25/1997 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

ND 


ARSENIC 

6020 

NA 

NA 


CADMIUM 

6020 

0.19 

0.22 


CALCIUM 

6010 

113 

103 


IRON 

6010 

0.37 

0.39 


LEAD 

6020 

NA 

NA 


MAGNESIUM 

6010 

58.8 

50.4 


MANGANESE 

6010 

4.9 

4.2 


NICKEL 

6010 

0.079 

0.065 


POTASSIUM 

6010 

3.4 J 

ND 


SILVER 

6020 

NA 

NA 


SODIUM 

6010 

18.3 

16.5 


ZINC 

6010 

95 

86 


ANIONS: 





SULFATE 

300.0 

571 

548 


SULFIDE TOTAL 

376.2 

NA 

NA 


FLUORIDE 

340.2 

NA 

NA 


CHLORIDE 

300.0 

17.5 

17.8 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

ND 

0.15 


NITRATE PLUSNITRITE ASN 

353.2 

0.11 

ND 


NITRITE AS N 

354.1 

0.002J 

0.0025J 


NITRATE ASN 

353.2/354.1 

NA 

NA 


AMMONIA 

350.1 

ND 

ND 


TOTAL SOLIDS: 





TSS 

160.2 

10.4 

14.8 


TDS 

160.1 

940 

869.0 


TOC 

9060 

0.71J 

1.8 


ALKALINITY, TOTAL: 





AS CaC03 

310.1 

84 

102 


ALKALINITY, BICARB 





ASCAC03 

310.1 

84 

102 


DISSOLVED OXYGEN (mg/L) 

— 

10.3 

7.5 


pH 

- 

7.2 

7.2 


CONDUCTIVITY (pS) 

- 

NA 

NA 


TEMPERATURE (degrees C) 

- 

9.2 

8.9 


* = Dissolved metals NA = Not analyzed 

— = Not applicable ND = Not detected 

pS = Microsiemens 
mg/L= Milligrams per liter 


69 
















Table A-2. Downflow Effluent Results 


DOWNFLOW EFFLUENT 





WED030994 

WED032394 

WED 040694 

WED042094 

WED050594 



ANALYTICAL 

03/09/94 

03/23/94 

04/06/94 

04/20/94 

05/05/94 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.021 

0.021 

0.027 

0.029 

0.033 


ARSENIC 

f ^ 

6020 

ND 

0.00056 

0.029 

0.016 

0.076 


CADMIUM 

6020 

0.00034 

0.00025 

0.00028 

0.00053 

0.00072 


CALCIUM 

6010 

105.0 

107.0 

110.0 

113.0 

113.0 


IRON 

6010 

1.5 

1.2 

1.1 

1.0 

1.1 


LEAD 

6020 

0.0015 

0.0012 

0.00065 

0.0015 

0.0017 


MAGNESIUM 

6010 

56.7 

56.9 

58.6 

58.3 

58.9 


MANGANESE 

6010 

1.6 

1.5 

1.5 

1.4 

1.4 


NICKEL 

6010 

0.0073 

0.0081 

0.0086 

0.010 

0.0090 


POTASSIUM 

6010 

55.8 

56.6 

54.0 

50.6 

48.3 


SILVER 

6020 

0.0015 

0.00012 

0.000060 

0.000089 

0.0051 


SODIUM 

6010 

19.0 

17.1 

18.1 

15.3 

18.6 


ZINC 

6010 

14.2 

14.9 

15.6 

15.3 

13.1 


ANIONS: 








SULFATE 

300.0 

350 

357 

338 

337 

280 


SULFIDE TOTAL 

376.2 

4.1 

5.2 

5.7 

2.1 

0.74 


FLUORIDE 

340.2 

0.82 

0.93 

0.88 

0.90 

0.87 


CHLORIDE 

300.0 

15.6 

28.4 

27.2 

28 

22 


PHOSPHORUS, TOTAL 

365.3 

9.9 

10.6 

11.0 

10.8 

10.4 


ORTHOPHOSPHATE 

365.3 

10.6 

12.4 

10.7 

11.1 

11.1 


NITRATE PLUS NITRITE AS N 

353.2 

0.24 

ND 

ND 

ND 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 


NITRATE AS N 

353.2/354.1 

0.24 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

5.4 

6.2 

5.9 

5.8 

4.6 


TOTAL SOLIDS: 

TSS 

160.2 

51.0 

27.0 

47.0 

39.2 

3.8 


TDS 

160.1 

864 

781 

766 

783 

753 


TOC 

9060 

60.4 

20.6 

29 

28.2 

20.8 


ALKALINITY, TOTAL: 

AS CaC03 

310.1 

193 

209 

200 

213 

193 


ALKALINITY, BICARB 

AS CAC03 

310.1 

193 

209 

200 

213 

193 


ORP (mV) 

— 

-77.0 


-180 


-184 


PH 

— 

7.3 


7.2 


7.6 


CONDUCTIVITY (pS) 

— 

845 


889 


803 


TEMPERATURE (degrees C) 

- 

4.1 


5.2 


8.8 


-- = Not applicable NA = Not analyzed 

pS = MicroSiemens ND = Not detected 

mgL = M illigrams per liter 
mV = M illivolts 


70 





















Table A-2 (continued). Downflow Effluent Results 


DOWNFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WED051994 

05/19/94 

mg/L 

VVED060194 

06/01/94 

mg/L 

YVED062994 

06/29/94 

mg/L 

WED071394 

07/13/94 

mg/L 

WED072894 

07/28/94 

mg/L 

WED081594 

08/15/94 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.024 

0.030 

0.017 

0.012 

0.017 

0.016 


ARSENIC 

6020 

0.066 

0.0013 

0.0011 

0.0010 

0.0012 

0.0011 


CADMIUM 

6020 

0.0011 

0.00073 

ND 

ND 

ND 

0.00033 


CALCIUM 

6010 

107.0 

112.0 

106.0 

118.0 

116.0 

114.0 


IRON 

6010 

1.0 

1.1 

1.0 

1.1 

1.1 

1.3 


LEAD 

6020 

0.0013 

0.0011 

ND 

ND 

ND 

ND 


MAGNESIUM 

6010 

57.1 

60.8 

55.2 

57.9 

55.9 

56.6 


MANGANESE 

6010 

1.3 

1.4 

1.5 

1.8 

1.8 

2.1 


NICKEL 

6010 

0.0088 

0.015 

0.014 

0.0089 

0.013 

0.013 


POTASSIUM 

6010 

39.5 

29.2 

19.8 

20.8 

17.8 

23.0 


SILVER 

6020 

0.000063 

ND 

0.00010 

0.00025 

ND 

0.00014 


SODIUM 

6010 

15.4 

15.2 

13.8 

14.7 

14.5 

15.5 


ZINC 

6010 

9.9 

10.3 

12.6 

15.3 

16.5 

14.5 


ANIONS: 









SULFATE 

300.0 

270 

319 

338 

337 

354 

311 


SULFIDE TOTAL 

376.2 

3.2 

2.4 

2.1 

1.3 

6.9 

1.5 


FLUORIDE 

340.2 

0.91 

0.95 

0.80 

0.90 

1.1 

1.0 


CHLORIDE 

300.0 

17.4 

18.4 

19.6 

17.8 

19.8 

19.2 


PHOSPHORUS, TOTAL 

365.3 

11.4 

10.1 

8.9 

9.5 

7.8 

8.7 


ORTHOPHOSPHATE 

365.3 

10.6 

9.2 

8.6 

8.6 

7.5 

6.7 


NITRATE PLUS NITRITE AS N 

353.2 

ND 

ND 

ND 

ND 

2.3 

1.7 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE AS N 

353.2/354.1 

ND 

ND 

ND 

ND 

2.3 

1.7 


AMMONIA 

350.1 

4.4 

3.2 

2.3 

3.1 

2.9 

3.2 


TOTAL SOLIDS: 









TSS 

160.2 

ND 

3.6 

33.6 

43 

45.6 

43.2 


TDS 

160.1 

739 

741 

709 

722 

747 

759 


TOC 

9060 

26.3 

35.6 

17.8 

15.9 

15.4 

15.6 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

196 

208 

188 

190 

188 

194 


ALKALINITY, BICARB 









AS CAC03 

310.1 

196 

208 

188 

190 

188 

194 


ORP (mV) 

— 

-271 


-253 

-250 

NA 

NA 


pH 

— 

7.28 


7.10 

7 

NA 

7.06 


CONDUCTIVITY (pS) 

- 

812 


1040 

1010 

996 

1006 

| TEM PERATURE (degrees C) 

- 

12.2 


12.3 

11.6 

11.8 

12.1 


- = Not applicable NA = Not analyzed 

pS = MicroSiemens ND = Not detected 

mgL = Milligrams per liter 
mV = M illivolts 


71 




















Table A-2 (continued). Downflow Effluent Results 




DOWN 

FLOW EF 

FLUENT 






ANALYTE 

ANALYTICAL 

METHOD 

WED082494 

08/24/94 

mg/L 

WED090794 

09/07/94 

mg/L 

WED091994 

09/19/94 

mg/L 

WED 100494 

10/04/94 

mg/L 

WED101994 

10/19/1994 

mg/L 

WED 110294 

11/02/94 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.015 

0.053 

0.022 

0.037 

0.018 

0.023 


ARSENIC 

6020 

0.0011 

ND 

0.0011 

0.0018 

ND 

ND 


CADMIUM 

6020 

0.00030 

ND 

ND 

0.00038 

0.00048 

0.00041 


CALCIUM 

6010 

117.0 

113.0 

124.0 

115.0 

112.0 

112.0 


IRON 

6010 

1.7 

1.8 

2.0 

1.8 

1.7 

1.8 


LEAD 

6020 

ND 

0.0016 

0.0023 

0.0032 

ND 

ND 


MAGNESIUM 

6010 

57.5 

55.8 

63.9 

57.6 

57.7 

58.0 


MANGANESE 

6010 

2.2 

2.0 

2.2 

1.9 

1.8 

1.6 


NICKEL 

6010 

0.014 

0.013 

0.020 

0.019 

0.020 

0.020 


POTASSIUM 

6010 

21.7 

25.0 

24.9 

21.6 

19.5 

16.8 


SILVER 

6020 

ND 

0.00032* 

0.00034 

0.0012 

ND 

ND 


SODIUM 

6010 

15.6 

14.5 

16.4 

14.4 

14.5 

15.5 


ZINC 

6010 

15.3 

15.2 

17.5 

15.5 

14.2 

12.1 


ANIONS: 

SULFATE 

300.0 

345 

349 

349 

333 

353 

365 


SULFIDE TOTAL 

376.2 

4.5 

0.12 

5.3 

10.7 

4.8 

7.4 


FLUORIDE 

340.2 

1.0 

0.94 

0.96 

0.88 

0.85 

0.87 


CHLORIDE 

300.0 

21.3 

22.3 

21.0 

21.0 

20.3 

20.8 


PHOSPHORUS, TOTAL 

365.3 

10.4 

1.6 

9.1 

8.8 

9.0 

8.2 


ORTHOPHOSPHATE 

365.3 

7.9 

8.6 

13.8 

8.5 

8.4 

8.8 


NITRATE PLUS NITRITE AS N 

353.2 

1.8 

ND 

ND 

ND 

ND 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE AS N 

353.2/354.1 

1.8 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

3.2 

2.6 

2.5 

2.9 

2.2 

1.5 


TOTAL SOLIDS: 

TSS 

160.2 

48.8 

49.6 

47.2 

52.0 

45.6 

40.0 


TDS 

160.1 

713 

741 

738 

716 

698 

734 


TOC 

9060 

13.8 

12.3 

10.3 

9.7 

8.1 

5.0 


ALKALINITY, TOTAL: 

AS CaC03 

310.1 

191 

194 

184 

200 

174 

152 


ALKALINITY, BICARB 

AS CAC03 

310.1 

191 

194 

184 

200 

174 

152 


ORP (mV) 

— 

-125 

-163 

-216 

-220 

-331 

-149 


pH 

— 

6.88 

6.91 

6.9 

6.9 

6.66 

6.92 


CONDUCTIVITY (pS) 

— 

973 

997 

1010 

960 

750 

890 


TEMPERATURE (degrees C) 

- 

13.4 

12.4 

10.7 

9.0 

6.8 

4.9 


-- = Not applicable NA = Not analyzed 

pS = M icroSiemens ND = Not detected 

mg/L = Milligrams per liter 
mV = M illivolts 


72 





















Table A-2 (continued). Downflow Effluent Results 


DOWNFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

MED 112094 

11/20/94 

mg/L 

MED 113094 

11/30/94 

mg/L 

MED121494 

12/14/94 

mg/L 

MED010495 

01/04/95 

mg/L 

MED011895 

01/18/95 

mg/L 

MED02019* 

02/01/95 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.018 

0.023 

0.013 

0.013 

0.014 

0.022 


ARSENIC 

6020 

ND 

ND 

ND 

0.0039 

0.0035 

ND 


CADMIUM 

6020 

0.00030 

0.00030 

0.00088 

ND 

ND 

ND 


CALCIUM 

6010 

120.0 

118.0 

120.0 

117.0 

119.0 

115.0 


IRON 

6010 

1.8 

2.4 

2.0 

2.7 

3.0 

2.6 


LEAD 

6020 

0.0054 

0.0018 

0.011 

ND 

0.0012 

ND 


MAGNESIUM 

6010 

60.6 

58.0 

56.6 

57.1 

54.5 

50.7 


MANGANESE 

6010 

1.6 

1.7 

1.5 

1.9 

1.9 

1.8 


NICKEL 

6010 

0.019 

0.019 

0.017 

0.013 

0.014 

0.018 


POTASSIUM 

6010 

16.0 

13.1 

11.5 

9.7 

9.9 

8.3 


SILVER 

6020 

ND 

0.00022 

ND 

ND 

ND 

ND 


SODIUM 

6010 

14.6 

14.5 

15.0 

14.3 

14.9 

15.0 


ZINC 

6010 

10.9 

11.7 

8.8 

8.3 

9.7 

10.5 


ANIONS: 









SULFATE 

300.0 

357 

391 

391 

386 

386 

380 


SULFIDE TOTAL 

376.2 

0.11 

5.8 

3.1 

3.3 

1.6 

2.3 


FLUORIDE 

340.2 

0.90 

1.1 

0.99 

1.1 

1.0 

1.0 


CHLORIDE 

300.0 

21.0 

22.0 

21.2 

22.1 

22.1 

21.9 


PHOSPHORUS, TOTAL 

365.3 

6.5 

7.2 

7.3 

6.6 

6.4 

6.3 


ORTHOPHOSPHATE 

365.3 

3.1 

5.0 

6.2 

5.5 

4.9 

6.0 


NITRATE PLUS NITRITE AS N 

353.2 

ND 

ND 

ND 

ND 

ND 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE AS N 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

2.2 

2.0 

0.41 

1.6 

1.5 

1.3 


TOTAL SOLIDS: 









TSS 

160.2 

41.0 

40.5 

28.5 

34.0 

37.0 

33.0 


TDS 

160.1 

750 

767 

744 

729 

718 

721 


TOC 

9060 

6.9 

20.4 

5.7 

4.8 

5.6 

4.8 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

187 

143 

152 

146 

141 

129 


ALKALINITY, BICARB 









ASCAC03 

310.1 

187 

143 

152 

146 

141 

129 


ORP (mV) 

— 

-170 

-220 

-195 

-20.0 

-6.5 

-7.3 


pH 

-- 

7.6 

7.12 

7.46 

7.26 

7.6 

7.6 


CONDUCTIVITY (pS) 

- 

NA 

600 

600 

590 

590 

670 


TEMPERATURE (degrees C) 

— 

3.7 

3.0 

2.9 

3.3 

3.0 

4.0 


-- = Not applicable NA = Not analyzed 

(j.S = M icroSiemens ND = Not detected 

mgE = Milligrams per liter 
mV = Millivolts 


73 



















Table A-2 (continued). Downflow Effluent Results 


DOWNFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WED021595 

02/15/95 

mg/L 

WED02279* 

02/27/95 

mg/L 

WED03159S 

03/15/95 

mg/L 

WED032995 

03/29/95 

mg/L 

WED041295 

04/12/95 

mg/L 

WED04269f 

04/26/95 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.018 

0.011 

0.011 

ND 

0.014 

ND* 


ARSENIC 

6020 

0.0011 

0.0019 

ND 

ND 

0.0021 

ND 


CADMIUM 

6020 

0.00033 

ND 

ND 

ND 

ND 

ND 


CALCIUM 

6010 

116.0 

121.0 

126.0 

103.0 

113.0 

109.0 


IRON 

6010 

2.4 

2.1 

2.2 

1.8 

1.8 

1.7 


LEAD 

6020 

0.0010 

ND 

ND 

ND 

ND 

ND 


MAGNESIUM 

6010 

51.2 

52.5 

54.3 

46.0 

48.1 

46.6 


MANGANESE 

6010 

1.9 

1.9 

2.1 

1.8 

1.9 

1.9 


NICKEL 

6010 

0.016 

0.015 

0.018 

0.019 

0.014 

0.014 


POTASSIUM 

6010 

8.4 

8.5 

9.0 

6.9 

6.7 

6.9 


SILVER 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


SODIUM 

6010 

15.9 

15.1 

16.5 

14.7 

14.1 

14.1 


ZINC 

6010 

10.7 

11.7 

13.0 

12.2 

12.6 

11.9 


ANIONS: 









SULFATE 

300.0 

359 

346 

370 

341 

338 

341 


SULFIDE TOTAL 

376.2 

1.9 

1.9 

3.1 

3.1 

0.099 

1.6 


FLUORIDE 

340.2 

1.1 

1.0 

1.1 

1.1 

1.0 

1.1 


CHLORIDE 

300.0 

22.1 

22.7 

24.4 

22.5 

21.8 

23.8 


PHOSPHORUS, TOTAL 

365.3 

17.5 

5.9 

5.7 

5.2 

4.7 

4.7 


ORTHOPHOSPHATE 

365.3 

5.1 

5.8 

5.7 

3.8 

5.4 

2.4 


NITRATE PLUS NITRITE AS N 

353.2 

ND 

ND 

ND 

ND 

ND 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE AS N 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

1.2 

1.2 

1.5 

1.3 

1.1 

1.1 


TOTAL SOLIDS: 









TSS 

160.2 

32.4 

34.0 

33.0 

31.0 

35.0 

31.2 


TDS 

160.1 

679 

723 

707 

662 

655 

651 


TOC 

9060 

4.3 

5.5 

5.4 

5.8 

6.9 

6.8 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

140 

152 

152 

141 

143 

141 


ALKALINITY, BICARB 









AS CAC03 

310.1 

140 

152 

152 

141 

143 

141 


ORP (mV) 

— 

59.0 

-82.0 

-65.0 

-81.1 

35.0 

NA 


pH 

- 

8.8 

7.1 

7.1 

7.3 

7.2 

NA 


CONDUCTIVITY (pS) 

- 

NA 

620 

680 

580 

580 

600 


TEM PERATURE (degrees C) 

- 

2.8 

5.6 

6.8 

5.6 

4.8 

7.0 


-- = Not applicable NA = Not analyzed 

pS = M icroSiemens ND = Not detected 

mg/L = Milligrams per liter 
mV = M illivolts 


74 






















Tabic A-2 (continued). Downflow Effluent Results 


DOWNFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WED051095 

05/10/95 

mg/L 

WED061295 

6/12/1995 

mg/L 

WED062895 

6/28/1995 

mg/L 

YVED071095 

7/10/1995 

mg/L 

VYED072695 

7/26/1995 

mg/L 

VYED08089f 

8/8/1995 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND* 

ND 

ND 

ND 

ND 

0.015 


ARSENIC 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CADMIUM 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CALCIUM 

6010 

121.0 

125 

142 

144 

157 

148 


IRON 

6010 

2.1 

4.2 

3.9 

3.9 

2.9 

2.8 


LEAD 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


MAGNESIUM 

6010 

47.8 

52.7 

61.9 

68.7 

71.7 

68.6 


MANGANESE 

6010 

2.4 

3.9 

4.4 

4.1 

4.1 

3.8 


NICKEL 

6010 

0.016 

0.017 

0.020 

0.021 

0.020 

0.022 


POTASSIUM 

6010 

6.5 

6.8 

7.1 

8.2 

7.6 

6.8 


SILVER 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


SODIUM 

6010 

14.1 

8.7 

10.6 

12.8 

12.6 

12.5 


ZINC 

6010 

13.3 

26.5 

31.2 

30.8 

29.7 

33.1 


ANIONS: 









SULFATE 

300.0 

348.0 

425 

453 

525 

537 

535 


SULFIDE TOTAL 

376.2 

0.38 

0.054 

6.9 

5.7 

0.83 

10.0 


FLUORIDE 

340.2 

1.1 

0.87 

0.80 

0.96 

0.86 

0.91 


CHLORIDE 

300.0 

22.6 

7.0 

7.2 

8.6 

10.1 

11.1 


PHOSPHORUS, TOTAL 

365.3 

4.3 

3.7 

4.7 

3.5 

2.6 

2.5 


ORTHOPHOSPHATE 

365.3 

4.1 

2.2 

1.5 

3.7 

2.0 

1.6 


NITRATE PLUS NITRITE AS N 

353.2 

ND 

ND 

ND 

ND 

ND 

ND 


NITRITE AS N 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

0.96 

0.90 

0.94 

1.0 

0.50 

0.64 


TOTAL SOLIDS: 









TSS 

160.2 

29.2 

43.0 

53.6 

48.0 

28.0 

38.8 


TDS 

160.1 

707 

763 

918 

946 

959 

1090 


TOC 

9060 

4.4 

6.6 

11.4 

5.4 

7.2 

4.7 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

137 

129 

195 

146 

141 



ALKALINITY, BICARB 



129 

195 





AS CAC03 

310.1 

137 



146 

141 



ORP (mV) 



-80 

-68 

-52 


14 


pH 



6.8 

6.6 

6.7 


7.1 


CONDUCTIVITY (pS) 



NA 

720 

NA 


850 


TEMPERATURE (degrees C) 



11.7 

12.3 

13.8 


14.1 


* - Aluminum was re-analyzed 6/2/95 due to blank contamination 
-- = Not applicable mV = Millivolts 

pS = M icroSiemens NA = Not analyzed 

mgd = M illigrams per liter ND = Not detected 


75 





















Tabic A-2 (continued). Downflow Effluent Results 


DOWNFLOW EFFLUENT 





WED082395 

VVED090595 

WED110995 

CDPHE 

CDPHE 

CDPHE 



ANALYTICAL 

8/23/1995 

9/5/1995 

11/9/1995 

1/29/1996 

2/29/1996 

4/25/1996 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

0.016 

ND 

NA 

NA 

NA 


ARSENIC 

6020 

ND 

ND 

ND 

NA 

NA 

NA 


CADMIUM 

6020 

ND 

ND 

0.00030 

0.00012 

0.00072 

0.15 


CALCIUM 

6010 

155 

147 

149 

NA 

NA 

NA 


IRON 

6010 

2.7 

2.2 

2.4 

NA 

0.28 

1.7 


LEAD 

6020 

ND 

ND 

0.0016 

NA 

NA 

NA 


MAGNESIUM 

6010 

70.2 

66.3 

66.2 

NA 

NA 

NA 


MANGANESE 

6010 

3.9 

3.7 

4.0 

3.2 

3.0 

2.2 


NICKEL 

6010 

0.026 

0.028 

0.04 

NA 

NA 

NA 


POTASSIUM 

6010 

6.2 

6.2 

5.3 

NA 

NA 

NA 


SILVER 

6020 

ND 

ND 

ND 

NA 

NA 

NA 


SODIUM 

6010 

13.7 

12.5 

14.5 

NA 

NA 

NA 


ZINC 

6010 

34.1 

29.1 

34.5 

28 

26 

15 


ANIONS: 

SULFATE 

300.0 

539 

529 

535 

440 

430 

318 


SULFIDE TOTAL 

376.2 

11.4 

5.6 

3.8 

NA 

NA 

NA 


FLUORIDE 

340.2 

0.85 

0.82 

0.81 

NA 

NA 

NA 


CHLORIDE 

300.0 

12.2 

14 

17.3 

NA 

NA 

NA 


PHOSPHORUS, TOTAL 

365.3 

3.0 

2.8 

2.5 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

3.0 

1.3 

1.1 

NA 

NA 

NA 


NITRATE PLUS NITRITE AS N 

353.2 

ND 

ND 

ND 

NA 

NA 

NA 


NITRITE AS N 

354.1 

0.0070 

ND 

ND 

NA 

NA 

NA 


NITRATE AS N 

353.2/354.1 

ND 

ND 

ND 

NA 

NA 

NA 


AMMONIA 

350.1 

0.78 

0.64 

0.39 

1.0 

1.1 

1.1 


TOTAL SOLIDS: 

TSS 

160.2 

50.0 

45.6 

12.8 

NA 

NA 

NA 


TDS 

160.1 

996 

941 

957 

NA 

NA 

NA 


TOC 

9060 

4.2 

4.9 

4.2 

NA 

NA 

NA 


ALKALINITY, TOTAL: 

AS CaC03 

310.1 

143 

179 

152 

NA 

NA 

NA 


ALKALINITY, BICARB 

AS CAC03 

310.1 

143 

179 

152 

NA 

NA 

NA 


ORP (mV) 




-60 

NA 

NA 

NA 


pH 




6.7 

NA 

NA 

NA 


CONDUCTIVITY (pS) 




750 

NA 

NA 

NA 


TEMPERATURE (degrees C) 




4.7 

NA 

NA 

NA 


— = Not applicable NA = Not analyzed 

pS = M icroSiemens ND = Not detected 

mg/L = Milligrams per liter 
mV = M illivolts 


76 



















Table A-2 (continued). Downflow Effluent Results 


DOWNFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

CDPHE 

5/31/1996 

mg/L 

CDPHE 

6/14/1996 

mg/L 

CDPHE 

7/19/1996 

mg/L 

CDPHE 

8/31/1996 

mg/L 

WED012197 

1/21/1997 

mg/1 

WED022097 

2/20/1997 

mg/1 

AQUEOUS 

ALUMINUM 

6010 

NA 

NA 

NA 

NA 

0.098 

ND 


ARSENIC 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


CADMIUM 

6020 

0.00016 

ND 

0.00021 

0.00013 

0.016 

0.034 


CALCIUM 

6010 

NA 

NA 

NA 

NA 

115 

113 


IRON 

6010 

0.87 

0.92 

1.10 

1.60 

0.53 

0.72 


LEAD 

6020 

NA 

NA 

NA 

NA 

57.3 

56.9 


MAGNESIUM 

6010 

NA 

NA 

NA 

NA 

3.3 

5.0 


MANGANESE 

6010 

1.8 

2.00 

2.10 

2.20 

NA 

NA 


NICKEL 

6010 

NA 

NA 

NA 

NA 

0.05 

0.035 


POTASSRJM 

6010 

NA 

NA 

NA 

NA 

0.39 

3.80 


SILVER 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


SODIUM 

6010 

NA 

NA 

NA 

NA 

16.6 

16 


ZINC 

6010 

11 

9.7 

8.7 

5.8 

55 

59.7 


ANIONS: 









SULFATE 

300.0 

230 

82 

340 

350 

421 

322 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

0.13 

ND 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

NA 

NA 

NA 

NA 

18.6 

18.6 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

NA 

NA 

NA 

NA 

1.1 

0.54 


NITRATE PLUS NITRITE AS N 

353.2 

NA 

NA 

NA 

NA 

ND 

0.2 


NITRITE AS N 

354.1 

NA 

NA 

NA 

NA 

0.0025 

0.0055 


NITRATE AS N 

353.2/354.1 

NA 

NA 

NA 

NA 

NA 

NA 


AMMONIA 

350.1 

0.67 

1.2 

0.90 

ND 

0.24 

0.20 


TOTAL SOLIDS: 









TSS 

160.2 

NA 

NA 

NA 

NA 

7.2 

6.0 


TDS 

160.1 

NA 

NA 

NA 

NA 

787 

752 


TOC 

9060 

NA 

NA 

NA 

NA 

7.2 

25.8 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

NA 

NA 

NA 

NA 

158 

259 


ALKALINITY, BICARB 









AS CAC03 

310.1 

NA 

NA 

NA 

NA 

158 

259 


ORP(mV) 


NA 

NA 

NA 

NA 

110 

92.0 


pH 


NA 

NA 

NA 

NA 

5.3 

7.0 


CONDUCTIVITY (uS) 


NA 

NA 

NA 

NA 

NA 

NA 


TEMPERATURE (degrees C) 


NA 

NA 

NA 

NA 

1.8 

1.8 


-- = Not applicable NA = Not analyzed 

|iS = M icroSiemens ND = Not detected 

mg/L = Milligrams per liter 
mV= Millivolts 


77 



















Table A-3. Upflow Effluent Results 


UPFLOW EFFLUENT 




WEU030994 

WEU032394 

WEU040694 

WEU042094 

WEU050594 

YVEU051994 


ANALYTICAL 

03/09/94 

03/23/94 

04/06/94 

04/20/94 

05/05/94 

05/19/94 

ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

ALUMINUM 

6010 

0.077 

0.20 

0.078 

0.39 

0.062 

0.028 

ARSENIC 

6020 

0.0062 

0.0071 

0.036 

0.028 

0.085 

0.067 

CADMIUM 

6020 

0.00042 

0.00049 

0.00034 

0.00036 

0.00024 

0.00020 

CALCIUM 

6010 

75.3 

96.2 

112.0 

115.0 

123.0 

115.0 

IRON 

6010 

0.48 

0.61 

0.48 

0.99 

0.27 

0.25 

LEAD 

6020 

0.0042 

0.0030 

0.0038 

0.020 

0.0022 

0.0015 

MAGNESIUM 

6010 

72.7 

71.4 

69.3 

63.1 

66.0 

60.1 

MANGANESE 

6010 

0.051 

0.072 

0.065 

0.16 

0.17 

0.25 

NICKEL 

6010 

0.0054 

0.0071 

0.0095 

0.0086 

0.0086 

0.0086 

POTASSIUM 

6010 

223.0 

188.0 

150.0 

108.0 

91.2 

49.4 

SILVER 

6020 

0.0014 

0.00015 

0.000084 

0.00048 

0.000071 

0.000072 

SODIUM 

6010 

33.9 

31.2 

27.3 

21.8 

22 

16.8 

ZINC 

6010 

0.22 

0.22 

0.13 

0.43 

0.14 

0.32 

ANIONS 








SULFATE 

300.0 

354 

388 

364 

343 

292 

265 

SULFIDE TOTAL 

376.2 

0.38 

7.9 

9.4 

1.9 

0.47 

2.4 

FLUORIDE 

340.2 

0.30 

0.57 

0.62 

0.72 

0.71 

0.88 

CHLORIDE 

300.0 

83.2 

76.0 

59.7 

50.0 

35.5 

21.8 

PHOSPHORUS TOTAL 

365.3 

24.3 

23.2 

20.5 

20.8 

18.3 

17.6 

ORTHOPHOSPHATE 

365.3 

26.8 

26.7 

20.9 

20.6 

18.6 

15.9 

NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

0.060 

ND 

ND 

ND 

NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 

NITRATE ASN 

353.2/354.1 

ND 

ND 

0.060 

ND 

ND 

ND 

AMMONIA 

350.1 

23.8 

19.6 

15.0 

12.9 

10.5 

6.8 

TOTAL SOLIDS 








TSS 

160.2 

6 

12.0 

6.0 

25.2 

ND 

ND 

TDS 

160.1 

1390 

1200 

1110 

1010 

934 

804 

TOC 

9060 

264 

51.3 

60.0 

49.3 

35.6 

23.8 

ALKALINITY, TOTAL: 








AS CaC03 

310.1 

367 

347 

310 

308 

265 

230 

ALKALINITY, BICARB 








ASCAC03 

310.1 

367 

347 

310 

308 

265 

230 

ORP (mV) 

- 

-377 


-280 


-269 

-271 

PH 

- 

8 


7.85 


7.20 

7.84 

CONDUCTIVITY (pS) 

- 

1410 


1222 


954 

893 

TEMPERATURE (degrees Q 

- 

5 


6.0 


7.8 

8.8 


AQUEOUS 


— = Not applicable NA = Not analyzed 

p/s = MicroSemens ND = Not detected 

mgT = Milligrams per liter 
mV= Millivolts 


78 





















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WEU060194 

06/01/94 

mg/L 

WEU062994 

06/29/94 

mg/L 

WEU071394 

07/13/95 

mg/L 

WEU072894 

07/28/95 

mg/L 

WEU081594 

08/15/94 

mg/L 

WEU082494 

08/24/94 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.045 

0.021 

ND 

0.38 

0.015 

0.023 


ARSENIC 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CADMIUM 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CALCIUM 

6010 

117 

120 

132 

132 

134 

132 


IRON 

6010 

0.26 

0.47 

0.79 

1.4 

2.7 

3.3 


LEAD 

6020 

0.0030 

0.0017 

ND 

ND 

ND 

ND 


MAGNESIUM 

6010 

61.5 

61.7 

61.4 

58.6 

58.3 

57.1 


MANGANESE 

6010 

0.33 

0.79 

1.3 

1.7 

2.1 

2.3 


NICKEL 

6010 

0.014 

0.011 

0.0052 

0.0075 

0.0089 

0.0077 


POTASSIUM 

6010 

37.3 

24.2 

17.3 

13.7 

12.8 

11.3 


SILVER 

6020 

ND 

0.00014 

0.00015 

ND 

0.00021 

ND 


SODIUM 

6010 

15.7 

15.6 

15 

14.2 

14.4 

14.4 


ZINC 

6010 

0.20 

0.35 

0.18 

0.29 

0.38 

0.58 


ANIONS 









SULFATE 

300.0 

330 

355 

372 

356 

369 

392 


SULFIDE TOTAL 

376.2 

5 

3.2 

0.59 

1.5 

0.69 

1.0 


FLUORIDE 

340.2 

0.81 

0.90 

0.80 

1.0 

0.96 

1.1 


CHLORIDE 

300.0 

22.2 

20.9 

18.9 

20.2 

19.9 

20.5 


PHOSPHORUS TOTAL 

365.3 

27.3 

12.8 

13.3 

10.8 

10.5 

9.8 


ORTHOPHOSPHATE 

365.3 

14.9 

21.3 

19.5 

10.5 

7.8 

9.2 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

ND 

1.9 

1.7 

1.8 


NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

0.077 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

1.9 

1.7 

1.8 


AMMONIA 

350.1 

5.6 

3.0 

3.0 

2.6 

1.6 

1.3 


TOTAL SOLIDS 









TSS 

160.2 

ND 

2.4 

2.0 

18.8 

7.6 

27.2 


TDS 

160.1 

808 

759 

766 

816 

802 

767 


TOC 

9060 

28.0 

11.4 

9.0 

9.6 

8.8 

6.0 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

244 

220 

211 

206 

194 

183 


ALKALINITY, BICARB 









ASCAC03 

310.1 

244 

220 

211 

206 

194 

183 


ORP (mV) 

— 


-275 

-280 

NA 

NA 

-344 


PH 

- 


7.7 

7.6 

NA 

7.6 

7.46 


CONDUCTIVITY (pS) 

- 


1115 

1090 

1049 

1069 

1037 


TEMPERATURE (degrees Q 

- 


9.7 

9.4 

9.7 

9.4 

10.0 


- = Not applicable NA = Not analyzed 

p/s = MicroSemens ND = Not detected 

mg/L = Milligrams per liter 
mV= Millivolts 


79 





















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 





WEU090794 

k'EU090794E 

WEU091994 

WEU100494 

WB5100494 

WEU101994 



ANALYTICAL 

09/07/94 

09/07/94 

09/19/94 

10/04/94 

10/04/94 

10/19/94 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.052 

ND 

0.023 

0.017 

6.1 

0.015 


ARSENIC 

6020 

ND 

ND 

ND 

0.0011 

0.0021 

0.0011 


CADMIUM 

6020 

ND 

ND 

ND 

ND 

0.024 

ND 


CALCIUM 

6010 

126 

0.15 

132 

127 

344 

128 


IRON 

6010 

4.3 

ND 

5.1 

5.7 

92.7 

5.6 


LEAD 

6020 

0.0011 

ND 

0.0015 

ND 

0.020 

ND 


MAGNESIUM 

6010 

53.3 

ND 

56.6 

54.5 

139.0 

54.1 


MANGANESE 

6010 

2.4 

ND 

2.6 

2.4 

28.6 

2.7 


NICKEL 

6010 

0.0083 

ND 

0.015 

0.015 

0.20 

0.019 


POTASSIUM 

6010 

10.2 

ND 

9.0 

11.9 

7.4 

7.7 


SILVER 

6020 

0.00011* 

ND* 

0.00046 

0.00052 

0.00099 

ND 


SODIUM 

6010 

13.6 

ND 

14.2 

13.8 

46.8 

14.6 


ZINC 

6010 

0.82 

0.019 

1.4 

2.4 

9.4 

3.1 


ANIONS 









SULFATE 

300.0 

395 

ND 

391 

369 

1760 

392 


SULFIDE TOTAL 

376.2 

0.12 

ND 

0.23 

5.0 

NS 

1.3 


FLUORIDE 

340.2 

1.0 

ND 

1.1 

0.99 

1.0 

0.95 


CHLORIDE 

300.0 

21.1 

ND 

20.4 

21.4 

6.0 

20.2 


PHOSPHORUS TOTAL 

365.3 

1.6 

ND 

7.9 

8.0 

NS 

6.8 


ORTHOPHOSPHATE 

365.3 

8.8 

ND 

9.8 

7.1 

ND 

6.8 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

ND 

ND 

NS 

ND 


NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

ND 

0.018 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

NS 

ND 


AMMONIA 

350.1 

1.0 

ND 

0.87 

1.0 

NS 

0.51 


TOTAL SOLIDS 









TSS 

160.2 

27.6 

ND 

28.8 

37.6 

49.6 

40.8 


TDS 

160.1 

787 

ND 

790 

750 

2520 

734 


TOC 

9060 

6.4 

ND 

5.8 

7.4 

NS 

5.3 


ALKALINITY, TOTAL: 

AS CaC03 

ALKALINITY, BICARB 

310.1 

175 

ND 

164 

182 

ND 

150 


ASCAC03 

310.1 

175 

ND 

164 

182 

ND 

150 


ORP (mV) 

- 

-315 


-267 

-260 

NA 

-344 


pH 

-- 

7.39 


7.3 

7.3 

5.2 

6.95 


CONDUCTIVITY (pS) 

-- 

1007 


990 

960 

NA 

760 


TEMPERATURE (degrees C) 

-- 

9.3 


9.2 

8.7 

15.0 

7.7 


— = Not applicable NA = Not analyzed 

|Vs = MicroSiemens ND = Not detected 

mgT. = Milligrams per liter 
mV= Millivolts 


80 





















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

VEU101994I 

10/19/94 

mg/L 

WEU110294 

11/02/94 

mg/L 

WEU112094 

11/20/94 

mg/L 

WEU113094 

11/30/94 

mg/1 

WEU121494 

12/14/94 

mg/L 

WEU010495 

01/04/95 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.025 

0.025 

0.012 

0.013 

0.015 

0.020 


ARSENIC 

6020 

ND 

0.0011 

0.0011 

0.0014 

0.0010 

0.0035 


CADMIUM 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CALCIUM 

6010 

130 

122 

127 

123 

127 

116 


IRON 

6010 

5.7 

7.0 

6.0 

7.5 

6.8 

6.3 


LEAD 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


MAGNESIUM 

6010 

54.8 

52.5 

53.8 

51.4 

52.4 

53.1 


MANGANESE 

6010 

2.7 

2.7 

2.8 

2.9 

2.9 

2.7 


NICKEL 

6010 

0.018 

0.018 

0.016 

0.018 

0.016 

0.012 


POTASSIUM 

6010 

7.6 

11.6 

9.6 

7.6 

7.6 

15.3 


SILVER 

6020 

ND 

ND 

0.00082 

0.00028 

ND 

0.00033 


SODIUM 

6010 

15 

14.2 

14.7 

14.5 

15.6 

14.5 


ZINC 

6010 

3.2 

6.8 

6.5 

7.9 

9.0 

11.7 


ANIONS; 









SULFATE 

300.0 

380 

371 

360 

379 

375 

341 


SULFIDE TOTAL 

376.2 

1.8 

3.8 

3.8 

4.6 

3.2 

3.3 


FLUORIDE 

340.2 

0.97 

1.1 

1.0 

1.2 

1.1 

1.1 


CHLORIDE 

300.0 

20.0 

23.2 

23.0 

22.2 

22.4 

25.6 


PHOSPHORUS, TOTAL 

365.3 

6.9 

6.2 

5.5 

6.9 

5.3 

4.8 


ORTHOPHOSPHATE 

365.3 

6.2 

5.9 

2.7 

2.7 

4.7 

3.0 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

ND 

ND 

ND 

ND 


NITRITE ASN 

354.1 

0.017 

0.016 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

0.52 

0.38 

0.74 

0.55 

1.5 

0.68 


TOTAL SOLIDS: 









TSS 

160.2 

36.8 

52.0 

49.0 

47.0 

44.0 

51.0 


TDS 

160.1 

742 

727 

745 

729 

729 

707 


TOC 

9060 

5.6 

9.4 

7.3 

19.1 

6.4 

12.5 


ALKALINITY, TOTAL: 









ASCaC03 

310.1 

148 

141 

185 

142 

157 

171 


ALKALINITY, BICARB 









ASCAC03 

310.1 

148 

141 

185 

142 

157 

171 


ORP (mV) 

- 

-344 

-164 

-160 

-216 

-196 

-80 


pH 

- 

6.95 

7.01 

7.2 

6.8 

7.33 

7.0 


CONDUCTIVITY (pS) 

-- 

760 

935 

NA 

640 

670 

670 


TEMPERATURE (degrees C) 

- 

7.7 

8.5 

8.1 

7.1 

7.7 

7.0 


— = Not applicable NA - Not applicable 

ps = MicroSemens ND - Not detected 

mg/L = Milligrams per liter 
mV= Millivolts 


81 




















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 





WEU011895 

WEU020195 

WEU021595 

WEU022795 

WEU031595 

WEU032995 



ANALYTICAL 

01/18/95 

02/01/95 

02/15/95 

02/27/95 

03/15/95 

03/29/95 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

0.026 

0.017 

0.012 

0.014 

0.012 

0.015 


ARSENIC 

6020 

0.0043 

0.0015 

0.0020 

0.0021 

0.0012 

0.0012 


CADMIUM 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


CALCIUM 

6010 

116 

119 

119 

116 

116 

105 


IRON 

6010 

5.4 

4.9 

4.3 

4.0 

4.0 

3.5 


LEAD 

6020 

0.0034 

ND 

ND 

ND 

ND 

ND 


MAGNESIUM 

6010 

49.5 

49.0 

49.1 

48.2 

48.2 

44.6 


MANGANESE 

6010 

2.6 

2.5 

2.5 

2.4 

2.4 

2.2 


NICKEL 

6010 

0.012 

0.016 

0.016 

0.016 

0.019 

0.019 


POTASSIUM 

6010 

10.5 

9.1 

9.1 

8.9 

7.5 

5.9 


SILVER 

6020 

ND 

ND 

Nr 

ND 

ND 

ND 


SODIUM 

6010 

15 

16.7 

16 

15.2 

16.0 

15.7 


ZINC 

6010 

12.5 

16.9 

12.9 

17.8 

18.0 

17.5 


ANIONS: 









SULFATE 

300.0 

347 

330 

308 

340 

335 

317 


SULFIDE TOTAL 

376.2 

3.0 

6.0 

3.3 

4.3 

2.7 

4.3 


FLUORIDE 

340.2 

1.1 

1.0 

1.1 

1.0 

1.1 

1.1 


CHLORIDE 

300.0 

23.0 

23.4 

23.4 

23.6 

24.3 

23.0 


PHOSPHORUS, TOTAL 

365.3 

4.7 

5.5 

13.3 

3.4 

4.1 

3.4 


ORTHOPHOSPHATE 

365.3 

3.0 

5.0 

3.7 

3.0 

2.6 

1.8 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

1.4 

ND 

ND 

ND 

ND 


NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

1.4 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

0.63 

0.52 

0.51 

0.34 

0.38 

0.31 


TOTAL SOLIDS: 









TSS 

160.2 

51.0 

54.0 

45.2 

47.0 

39.0 

41.0 


TDS 

160.1 

693 

692 

682 

700 

671 

667 


TOC 

9060 

7.8 

6.8 

6.2 

5.8 

4.3 

7.0 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

168 

161 

191 

150 

151 

154 


ALKALINITY. BICARB 









AS CAC03 

310.1 

168 

161 

191 

150 

151 

154 


ORP (mV) 

- 

5 

-11.7 

-44.0 

-65 

-63 

-81.1 


pH 

- 

7.1 

7.4 

7.2 

6.9 

6.9 

7.3 


CONDUCTIVITY (pS) 

-- 

650 

610 

NA 

680 

650 

580 


TEMPERATURE (degrees C) 

-- 

8.3 

6.1 

7.6 

8.4 

8.8 

5.6 


— = Not applicable NA = Not analyzed 

pS = MicroSiemens ND = Not detected 

mgl. = Milligrams per liter 
mV = Millivolts 


82 





















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WEU041295 

04/12/95 

mg/L 

WEU042695 

04/26/95 

mg/L 

WEU051095 

05/10/95 

mg/L 

WEU061295 

06/12/95 

mg/L 

WEU062895 

6/28/1995 

mg/L 

WEU071095 

7/10/1995 

mg/L 

AQ UEOUS 

ALUMINUM 

6010 

0.013 

ND* 

ND* 

0.028 

ND 

ND 


ARSENIC 

6020 

0.0028 

ND 

ND 

ND 

ND 

ND 


CADMIUM 

6020 

ND 

0.00078 

0.0094 

0.0084 

0.0045 

ND 


CALCIUM 

6010 

114 

106 

110 

103 

121 

130 


IRON 

6010 

3.5 

2.2 

2.2 

4.6 

3.7 

3.8 


LEAD 

6020 

ND 

ND 

0.0019 

0.0018 

ND 

ND 


MAGNESIUM 

6010 

46.5 

45.3 

44.5 

45.2 

60.2 

68.2 


MANGANESE 

6010 

2.5 

2.0 

2.5 

3 

4.0 

4.1 


NICKEL 

6010 

0.013 

0.015 

0.022 

0.019 

0.026 

0.026 


POTASSIUM 

6010 

7.1 

11.7 

18.1 

7.5 

6.0 

5.4 


SILVER 

6020 

ND 

ND 

ND 

ND 

ND 

ND 


SODIUM 

6010 

15.3 

14.2 

13.3 

8.9 

11.2 

13.2 


ZINC 

6010 

15.9 

18.5 

26.7 

33.5 

47.1 

50.8 


ANIONS: 









SULFATE 

300.0 

326 

326 

355 

326 

494 

514 


SULFIDE TOTAL 

376.2 

0.39 

2.9 

1.3 

0.065 

1.5 

1.5 


FLUORIDE 

340.2 

1.0 

1.1 

1.2 

0.90 

0.90 

0.96 


CHLORIDE 

300.0 

22.5 

26.0 

25.9 

7.6 

7.0 

8.3 


PHOSPHORUS, TOTAL 

365.3 

3.0 

3.2 

2.0 

2.3 

1.2 

1.5 


ORTHOPHOSPHATE 

365.3 

2.7 

1.6 

2.4 

1.5 

0.34 

0.48 


NITRATE PLUS NITRITE ASN 

353.2 

ND 

ND 

ND 

ND 

ND 

ND 


NITRITE ASN 

354.1 

ND 

ND 

ND 

ND 

ND 

ND 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

ND 


AMMONIA 

350.1 

0.33 

0.31 

0.42 

0.36 

0.20 

0.20 


TOTAL SOLIDS: 









TSS 

160.2 

41.9 

31.2 

29.0 

47.3 

18.8 

25.6 


TDS 

160.1 

657 

607 

724 

668 

885 

944 


TOC 

9060 

8.0 

8.3 

9.9 

9.1 

4.9 

4.8 


ALKALINITY, TOTAL: 









ASCaC03 

310.1 

152 

147 

138 

181 

136 

147 


ALKALINITY, BICARB 









ASCAC03 

310.1 

152 

147 

138 

181 

136 

147 


ORP (mV) 

— 

-7.0 

NA 


-57 




pH 

— 

7.1 

NA 


6.7 

6.9 

6.9 


CONDUCTIVITY (pS) 

- 

620 

620 


NA 




TEMPERATURE (degrees C) 

- 

8.5 

8.0 


10.1 




* - Aluminum was re-analyzed 6/2/95 due to blank contamination 
— = Not applicable Not analyzed 

jjS = MicroSiemens ND = Not detected 

mg/L = Milligrams per liter 
mV = Millivolts 


83 



















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 





WEU072695 

VVEU080895 

WEU082395 

WEU090595 

WEU110995 

CDPHE 



ANALYTICAL 

7/26/1995 

8/8/1995 

8/23/1995 

9/5/1995 

11/9/1995 

1/29/1996 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

ND 

ND 

ND 

ND 

NA 


ARSENIC 

6020 

ND 

ND 

ND 

ND 

ND 

NA 


CADMIUM 

6020 

0.0060 

0.0046 

0.0093 

0.010 

0.04400 

0.037 


CALCIUM 

6010 

144 

135 

141 

137 

133 

NA 


IRON 

6010 

2.5 

2.5 

2.1 

1.8 

0.93 

1.6 


LEAD 

6020 

ND 

ND 

ND 

ND 

0.0022 

NA 


MAGNESIUM 

6010 

68.6 

64.4 

66.1 

64.3 

62.1 

NA 


MANGANESE 

6010 

4.1 

3.8 

3.8 

3.6 

4.4 

3.3 


NICKEL 

6010 

0.028 

0.032 

0.036 

0.04 

0.059 

NA 


POTASSIUM 

6010 

5.7 

4.9 

4.5 

ND 

4.3 

NA 


SILVER 

6020 

ND 

ND 

ND 

ND 

ND 

NA 


SODIUM 

6010 

12.2 

13.3 

14.0 

12.2 

15.6 

NA 


ZINC 

6010 

53.2 

56.6 

59.8 

59.9 

73.6 

47 


ANIONS: 









SULFATE 

300.0 

549 

584 

561 

569 

559 

460 


SULFIDE TOTAL 

376.2 

4.3 

3.5 

5.2 

2.8 

0.84 

NA 


FLUORIDE 

340.2 

0.89 

0.88 

0.90 

0.86 

0.96 

NA 


CHLORIDE 

300.0 

10.0 

11.2 

12.5 

13.7 

17.1 

NA 


PHOSPHORUS, TOTAL 

365.3 

1.1 

1.1 

1.3 

1.5 

0.69 

NA 


ORTHOPHOSPHATE 

365.3 

1.0 

0.9 

1.2 

0.43 

0.80 

NA 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

ND 

ND 

ND 

NA 


NITRITE ASN 

354.1 

ND 

ND 

0.0080 

ND 

ND 

NA 


NITRATE ASN 

353.2/354.1 

ND 

ND 

ND 

ND 

ND 

NA 


AMMONIA 

350.1 

0.11 

ND 

0.21 

ND 

ND 

0.2 


TOTAL SOLIDS: 









TSS 

160.2 

16.8 

17.6 

30.0 

26 

5.2 

NA 


TDS 

160.1 

961 

999 

1010 

978 

932 

NA 


TOC 

9060 

5.7 

3.4 

3.2 

3.7 

2.1 

NA 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

135 

138 

149 

160 

115 

NA 


ALKALINITY, BICARB 









AS CAC03 

310.1 

135 

138 

149 

160 

115 

NA 


ORP (mV) 

— 








PH 

— 

6.8 

7.1 

7.1 

6.9 

7.0 

NA 


CONDUCTIVITY (pS) 

-- 






NA 


TEMPERATURE (degrees C) 

— 






NA 


— = Not applicable NA = Not analyzed 

pS = MicroSemens ND = Not detected 

mg/L = Milligrams per liter 
mV = Millivolts 


84 




















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 





CDPHE 

CDPHE 

CDPHE 

CDPHE 

CDPHE 

CDPHE 



ANALYTICAL 

2/29/1996 

4/25/1996 

5/31/1996 

6/14/1996 

7/19/1996 

8/31/1996 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

NA 

NA 

NA 

NA 

NA 

NA 


ARSENIC 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


CADMIUM 

6020 

0.035 

0.030 

0.140 

0.031 

0.051 

0.053 


CALCIUM 

6010 

NA 

NA 

NA 

NA 

NA 

NA 


IRON 

6010 

1.3 

0.81 

0.17 

1.1 

0.87 

0.90 


LEAD 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


MAGNESIUM 

6010 

NA 

NA 

NA 

NA 

NA 

NA 


MANGANESE 

6010 

3.1 

2.3 

2.7 

2.2 

2.6 

2.5 


NICKEL 

6010 

NA 

NA 

NA 

NA 

NA 

NA 


POTASSIUM 

6010 

NA 

NA 

NA 

NA 

NA 

NA 


SILVER 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


SODIUM 

6010 

NA 

NA 

NA 

NA 

NA 

NA 


ZINC 

6010 

42 

31 

56 

30 

41 

43.0 


ANIONS: 









SULFATE 

300.0 

430 

329 

420 

310 

410 

45 


SULFIDE TOTAL 

376.2 

NA 

NA 

NA 

NA 

NA 

NA 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

NA 

NA 

NA 

NA 

NA 

NA 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

NA 

NA 

NA 

NA 

NA 

NA 


NITRATE PLUSNITRITE ASN 

353.2 

NA 

NA 

NA 

NA 

NA 

NA 


NITRITE ASN 

354.1 

NA 

NA 

NA 

NA 

NA 

NA 


NITRATE ASN 

353.2/354.1 

NA 

NA 

NA 

NA 

NA 

NA 


AMMONIA 

350.1 

0.4 

0.3 

ND 

0.2 

0.2 

ND 


TOTAL SOLIDS: 









TSS 

160.2 

NA 

NA 

NA 

NA 

NA 

NA 


TDS 

160.1 

NA 

NA 

NA 

NA 

NA 

NA 


TOC 

9060 

NA 

NA 

NA 

NA 

NA 

NA 


ALKALINITY, TOTAL: 









ASCaC03 

310.1 

NA 

NA 

NA 

NA 

NA 

NA 


ALKALINITY, BICARB 









ASCAC03 

310.1 

NA 

NA 

NA 

NA 

NA 

NA 


ORP (mV) 

— 

NA 

NA 

NA 

NA 

NA 

NA 


PH 

- 

NA 

NA 

NA 

NA 

NA 

NA 


CONDUCTIVITY (pS) 

— 

NA 

NA 

NA 

NA 

NA 

NA 


TEMPERATURE (degrees C) 

-- 

NA 

NA 

NA 

NA 

NA 

NA 


— = Not applicable NA = Not analyzed 

pS = MicroSiemens ND = Not detected 

mgT, = Milligrams per liter 
mV = Millivolts 


85 





















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 





WEU120996 

YVEU012197 

WEU022097 

VVEU031997 

WEU042297 




ANALYTICAL 

12/09/96 

01/21/97 

02/20/96 

03/19/97 

04/22/97 

04/22/97 


ANALYTE 

METHOD 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

ND 

ND 

ND 

ND 

ND (D) 


ARSENIC 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


CADMIUM , 

6020 

0.088 

0.032 

0.057 

0.034 

0.015 

0.015 (D) 


CALCIUM 

6010 

115 

116 

119 

109 

95.9 

97.3 (D) 


IRON 

6010 

0.99 

1.2 

0.8 

1.1 

0.98 

0.99 (D) 


LEAD 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


MAGNESIUM 

6010 

53.4 

52.8 

55.4 

47.8 

43.2 

43.9 (D) 


MANGANESE 

6010 

2.9 

2.7 

3 

2.8 

2 

2(D) 


NICKEL 

6010 

0.035 J 

0.032 J 

0.033 J 

0.041 

0.021 J 

0.021 J (D) 


POTASSIUM 

6010 

3.6 J 

3.5 J 

4.4 

ND 

4.6 J 

4.6 J 


SILVER 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


SODIUM 

6010 

15.3 

15.5 

15.7 

14.4 

7.3 

7(D) 


ZINC 

6010 

46 

41.3 

48.6 

38 

22.7 

22.9 (D) 


ANIONS: 









SULFATE 

300.0 

434 

400 

413 

392 

252 

248 (D) 


SULFIDE TOTAL 

376.2 

0.8 

1.2 

0.71 

0.037 J 

3.5 

3.6(D) 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

17.4 

18.1 

19.4 

18.5 

12.4 

12.2 (D) 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

0.84 

1.7 

1.1 

2 

1.7 

1.8(D) 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 

ND 

ND 

0.020 J 

ND (D) 


NITRITE ASN 

354.1 

ND 

0.0057 J 

0.0055 J 

0.0058 J 

0.0038 J 

.0036 J(D) 


NITRATE ASN 

353.2/354.1 

NA 

NA 

NA 

NA 

NA 

NA 


AMMONIA 

350.1 

0.070 J 

0.17 

0.11 

0.19 

0.22 

0.21 (D) 


TOTAL SOLIDS 









TSS 

160.2 

NA 

17.2 

7.2 

16.4 

21.6 

23.2 (D) 


TDS 

160.1 

806 

773 

800 

712 

575 

574 (D) 


TOC 

9060 

5.6 

5.2 

5.3 

5 

7.6 

7.7 (D) 


ALKALINITY, TOTAL: 

ASCaC03 

310.1 

158 

178 

160 

153 

180 

179 (D) 


ALKALINITY, BICARB 

ASCAC03 

310.1 

158 

178 

160 

153 

180 

179 (D) 


ORP (mV) 

- 

94 

108 

80 

82 

72 

72 


pH 

- 

6.5 

5.4 

6.7 

6.3 

6.2 

6.2 


CONDUCTIVITY (pS) 

- 

NA 

NA 

NA 

NA 

NA 

NA 


TEMPERATURE (degrees C) 

- 

5.1 

3.2 

5.0 

6.0 

7.0 

7.0 


— = Not applicable NA = Not analyzed 

jjS = MicroSiemens ND = Not detected 

mg/1 = Milligrams per liter 
mV = Millivolts 


86 




















Table A-3 (continued). Upflow Effluent Results 


UPFLOW EFFLUENT 



ANALYTE 

ANALYTICAL 

METHOD 

WEU052897 

05/28/97 

mg/L 

WEU062397 

06/23/97 

mg/L 

WEU082897 

8/28/1997 

mg/L 

WEU093097 

9/30/1997 

mg/L 

WEU102997 

10/29/1997 

mg/L 

WEU112597 

11/25/1997 

mg/L 

AQUEOUS 

ALUMINUM 

6010 

ND 

ND 

0.10 

0.078 

ND 

ND 


ARSENIC 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


CADMIUM 

6020 

0.2 

ND 

0.0063 

0.0040 

0.010 

0.016 


CALCIUM 

6010 

99.6 

113 

153 

152 

144 

138 


IRON 

6010 

3.3 

1.2 

4.0 

2.9 

1.2 

1 


LEAD 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


MAGNESIUM 

6010 

48.9 

53.7 

64.6 

64.8 

65.6 

56.6 


MANGANESE 

6010 

2.1 

2.2 

3.0 

2.7 

3.6 

3.0 


NICKEL 

6010 

0.022 J 

ND 

0.023 

ND 

ND 

ND 


POTASSIUM 

6010 

4.0 J 

4.9 J 

5.2 

4.5 

3.9 

ND 


SILVER 

6020 

NA 

NA 

NA 

NA 

NA 

NA 


SODIUM 

6010 

14.8 

0.94 J 

6.9 

13.5 

14.7 

15.0 


ZINC 

6010 

60.1 

25.4 

21.2 

14.8 

26.4 

24.6 


ANIONS: 









SULFATE 

300.0 

250 

275 

308 

311 

484 

460 


SULFIDE TOTAL 

376.2 

17 

6.1 

2.4 

4.1 

2.2 

2.4 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

6.9 

9.2 

12.9 

15.5 

16.7 

17.8 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

1.1 

1.9 

0.70 

2.7 

1.1 

1.9 


NITRATE PLUSNITRITE ASN 

353.2 

ND 

ND 


0.034 

ND 

ND 


NITRITE ASN 

354.1 

0.0051 J 

0.0047J 

0.0040 

ND 

0.0020J 

0.0032J 


NITRATE ASN 

353.2/354.1 

NA 

NA 

NA 

NA 

NA 

NA 


AMMONIA 

350.1 

ND 

0.4 

1.7 

1.3 

0.64 

1.0 


TOTAL SOLIDS: 









TSS 

160.2 

102 

38 


31.6 

13.6 

27.2 


TDS 

160.1 

566 

683 

808 

865 

892 

887 


TOC 

9060 

5.3 

16 

29.7 

18.8 

6.7 

6.6 


ALKALINITY, TOTAL: 









AS CaC03 

310.1 

190 

228 

NA 

317 

166 

199 


ALKALINITY, BICARB 









ASCAC03 

310.1 

190 

228 

NA 

317 

166 

199 


ORP (mV) 

— 

-58 

47 

30 

-37 

NA 

49 


PH 

— 

6.4 

6.8 

5.7 

6.2 

6.7 

6.7 


CONDUCTIVITY (pS) 

- 

NA 

NA 

NA 

NA 

NA 

NA 


TEMPERATURE (degrees C) 

- 

9.2 

12.7 

12.0 

10.3 

5.7 

5.5 


— = Not applicable NA = Not analyzed 

pS = MicroSiemens ND = Not detected 

mg/L = Milligrams per liter 
mV = Millivolts 


87 




















Table A-4. Substrate Results - Downflow Cell 


SUBSTRATE - DOWNFLOW CE1 

LL 




SD2032394 

SD2062994 

SD5062994 

SD5082594 



ANALYTICAL 

03/23/94 

06/29/94 

06/29/94 

08/25/94 


ANALYTE 

METHOD 

mg/kg 

mg/kg 

mg/kg 

mg/kg 

SEDIMENT 

ALUMINUM 

6010 

1410.0 

65.6 

423.0 

2580.0 


ARSENIC 

6020 

2.9 

0.14 

ND 

0.59 


CADMIUM 

6020 

2.2 

0.56 

4.8 

5.1 


CALCIUM 

6010 

7040.0 

406 

2330.0 

7650.0 


IRON 

6010 

2250.0 

88.7 

653.0 

3650.0 


LEAD 

6020 

7.4 

3.1 

53.4 

16.2 


MAGNESIUM 

6010 

2140.0 

145 

571.0 

2120.0 


MANGANESE 

6010 

99.2 

4.1 

36.0 

140.0 


NICKEL 

6010 

3.9 

ND 

1.9 

4.9 


POTASSIUM 

6010 

890.0 

149.0 

184.0 

1360.0 


SILVER 

6020 

0.061 

0.024 

0.79 

0.16 


SODIUM 

6010 

ND 

76.3 

ND 

ND 


ZINC 

6010 

1560.0 

59.7 

1000.0 

2650.0 


ANIONS: 

SULFATE 

300.0 

214 

56.5 

143.0 

214 


SULFIDE, REACTIVE 

EPA/OSW 

0.40 

19.1 

18.6 

3.2 


SULFIDE, ACID VOLATILE 

EPA (Draft) 

NA 

226 

178.0 

ND 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

NA 

NA 

NA 

NA 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

25.8 

63.4 

30.5 

18.8 


NITRATE PLUS NITRITE AS N 

353.2 

NA 

NA 

NA 

NA 


NITRITE AS N 

354.1 

NA 

NA 

NA 

NA 


NITRATE AS N 

353.2/354.1 

NA 

NA 

NA 

NA 


AMMONIA 

350.1 

NA 

NA 

NA 

NA 


WATER (%) 

ILMOl.l 

82 

62 

70 

75 


NA = Not analyzed 
ND = Not detected 


88 
















Table A-4 (continued). Substrate Results - Downflow Cell 


SUBSTRATE - DOWN! 

FLOW CE1 

LL 




SD2100494 

SD5100494 

SD2110294 

SD2010495 



ANALYTICAL 

10/04/94 

10/04/94 

11/02/94 

01/04/95 


ANALYTE 

METHOD 

mg/kg 

mg/kg 

mg/kg 

mg/kg 

SEDIMENT 

ALUMINUM 

6010 

2640.0 

3200.0 

3200.0 

2430.0 


ARSENIC 

6020 

1.5 

0.97 

1.3 

1.5 


CADMIUM 

6020 

4.6 

10.5 

4.3 

4.3 


CALCIUM 

6010 

8460.0 

4890.0 

11700.0 

8770.0 


IRON 

6010 

3410.0 

4640.0 

4860.0 

3460.0 


LEAD 

6020 

46.4 

30.8 

11.3 

18.2 


MAGNESIUM 

6010 

2180.0 

1800.0 

2910.0 

2190.0 


MANGANESE 

6010 

160.0 

151.0 

232.0 

144.0 


NICKEL 

6010 

3.7 

6.4 

7.0 

4.9 


POTASSIUM 

6010 

930.0 

1410.0 

1140.0 

729.0 


SILVER 

6020 

0.17 

0.29 

0.069 

0.28 


SODIUM 

6010 

ND 

108.0 

92.8 

ND 


ZINC 

6010 

1510.0 

2850.0 

3170.0 

3250.0 


ANIONS: 







SULFATE 

300.0 

86.8 

187.0 

159.0 

184.0 


SULFIDE, REACTIVE 

EPA/OSW 

103.0 

79.3 

1.1 

15.3 


SULFIDE, ACID VOLATILE 

EPA (Draft) 

190.0 

70.6 

171.0 

117.0 


FLUORIDE 

340.2 

NA 

NA 

NA 

NA 


CHLORIDE 

300.0 

NA 

NA 

NA 

NA 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

39.0 

3.3 

12.6 

6.4 


NITRATE PLUS NITRITE AS N 

353.2 

NA 

NA 

NA 

NA 


NITRITE AS N 

354.1 

NA 

NA 

NA 

NA 


NITRATE AS N 

353.2/354.1 

NA 

NA 

NA 

NA 


AMMONIA 

350.1 

NA 

NA 

NA 

NA 


WATER (%) 

ILM01.0 

62 

70 

NA 

63 


NA = Not analyzed 
ND = Not detected 


89 















Table A-4 (continued). Substrate Results - Downflow Cell 


SUBSTRATE - DOWNFLOW CELL 



ANALYTE 

ANALYTICAL 

METHOD 

SD2061295 

06/12/95 

mg/kg 

SD2082395 

34934 

mg/kg 

SD093097 

09/30/97 

mg/kg 

SEDIMENT 

ALUMINUM 

6010 

2050 

1660 

2200 


ARSENIC 

6020 

0.59 

0.75 

NA 


CADMIUM 

6020 

11.3 

31.4 

219 


CALCIUM 

6010 

7860 

4720 

7680 


IRON 

6010 

3200 

2490 

4400 


LEAD 

6020 

21.4 

177 

NA 


MAGNESIUM 

6010 

1860 

1360 

2070 


MANGANESE 

6010 

149 

108 

1950 


NICKEL 

6010 

7.0 

6.2 

22.5 


POTASSIUM 

6010 

646 

463 

666 


SILVER 

6020 

0.11 

ND 

NA 


SODIUM 

6010 

119 

ND 

1930 


ZINC 

6010 

4990 

4680 

37500 


ANIONS: 






SULFATE 

300.0 

93.0 

154 

154 


SULFIDE, REACTIVE 

EPA/OSW 

5.3 

2.5 

NA 


SULFIDE, ACID VOLATILE 

EPA (Draft) 

528 

687 

187 


FLUORIDE 

340.2 

MNA 

NA 

NA 


CHLORIDE 

300.0 

NA 

NA 

NA 


PHOSPHORUS, TOTAL 

365.3 

NA 

NA 

NA 


ORTHOPHOSPHATE 

365.3 

5.0 

1.8 

NA 


NITRATE PLUS NITRITE AS N 

353.2 

NA 

NA 

NA 


NITRITE AS N 

354.1 

NA 

NA 

NA 


NITRATE AS N 

353.2/354.1 

NA 

NA 

NA 


AMMONIA 

350.1 

NA 

NA 

NA 


WATER (%) 

ILM01.0 


60 

64 


NA = Not analyzed 
ND = Not detected 


90 












Appendix B 
Case Study 


91 



BUREAU OF MINES 
INFORMATION CIRCULAR/1994 



PB94173341 
II III II Mill llll II llllll 


Passive Treatment of Coal Mine 
Drainage 


By Robert S. Hedln, Robert W. Narin, 
and Robert L. P. Kleinmann 


UNITED STATES DEPARTMENT OF THE INTERIOR 


REPHOOUCIO «Y: 

U.S. Otpartmcnt of Commorco 
National Tocfcncal Information Sorvtot 

(/Ma* fit 


92 






























































































































































































U.S. Department of the Interior 
Mission Statement 

As the Nation’s principal conservation agency, the Department of 
the Interior has responsibility for most of our nationally-owned 
public lands and natural resources. This includes fostering 
sound use of our land and water resources; protecting our fish, 
wildlife, and biological diversity; preserving the environmental 
and cultural values of our national parks and historical places; and 
providing for the enjoyment of life through outdoor recreation. 
The Department assesses our energy and mineral resources and 
works to ensure that their development is in the best interests of 
all our people by encouraging stewardship and citizen participa¬ 
tion in their care. The Department also has a major responsibility 
for American Indian reservation communities and for people who 
live in island territories under U.S. administration. 


X 


93 




information Circular 9389 


Passive Treatment of Coal Mine 
Drainage 


By Robert S. Hedin, Robert W. Nairn, 
and Robert L. P. Kleinmann 


UNITED STATES DEPARTMENT OF THE INTERIOR 
Bruce Babbitt, Secretary 


BUREAU OF MINES 


Library of Congress Cataloging in Publication Data: 


Hedin, Robert S., 1956- 

Passive treatment of coal mine drainage / by Robert S. Iledin, Robert W. Naim, 
and Robert L.P. Kieinmann. 

p. cm. — (Information circular; 9389) 

Includes bibliographical references (p. 34). 

Supt. of Docs, no.: I 28.27:9389. 

1. Mine drainage. 2. Coal mines and mining—Waste disposal. 3. Mine water— 
Purification. 4. Water—Purification—Biological treatment. I. Naim, Robert W. 
II. Kieinmann, Robert L P. III. Title. IV. Series: Information circular (United 
States. Bureau of Mines); 9389. 


TN295.U4 [TN321] 622 s-dc20 [622*_5J 93-23717 CIP 


7 /,' 


95 





REPORT DOCUMENTATION PAGE 

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^ III ill!!llllllllllllllllllllll!!!llll 

2. REPORT DATE 

3. REPORT TYPE ANO DATE* COVB*ED 

PB94-173341 

October 4, 1993 

Information Circular 9389 

4. TITL£ AND SUBTITLE 

Passive Treatment of Coal Mine Drainage 

5. FUNblNQ NUMBER* 

6. AUTHOmSI 

Hedin, R. S., Nairn, R. W., and Kleinmann, R. L. P. 

(. PERFORMING ORGANIZATION NAME(S) AND AODRt&SlESI 

U.S. Bureau of Mines 

Pittsburgh Research Center 

P.O. Box 18070 

Pittsburgh, PA 15236 

b. PERFORMING 

ORGANIZATION REPORT NUMBER 

IC 9389 

9. SPONSOWNG/MON1 TORINO AGENCY INAM&S) AND AODflE3*(Cil 

U.S. Bureau of Mines 

Research 

810 7th Street, NW 

Washington, DC 20241 

^0. SPONSORING/MONITOAlNQ 

AGENCY REPORT NUMBER 

SUPH-EMEnTARV NOTES 

None 

12A. DJSTRlAUTION/AVAl LABILITY STATEMENT 

12B. tXSlRJBUTION CODE 


Passive methods of treating mine water utilize chemical and biological processes that decrease metal concentrations and 
neutralize acidity. Compared to conventional chemical treatment, passive methods generally require more land area, but utilize less 
costly reagents and require less operational attention and maintenance. Currently, three types of passive technologies exist: 
aerobic wetlands, wetlands that contain an organic substrate, and anoxic limestone drains. Aerobic wetlands promote mixed 
oxidation and hydrolysis reactions, and are most effective when the raw mine water is net alkaline. Organic substrate wetlands 
promote anaerobic bacterial activity that results in the precipitation of metal sulfides and the generation of bicarbonate alkalinity. 
Anoxic limestone drains generate bicarbonate alkalinity and can be useful for the pretreatment of mine water before it flows into a 
wetland. 

Rates of metal and acidity removal for passive systems have been developed empirically. Aerobic wetlands remove Fe and 
Mn from alkaline water at rates of 10-20 g • m' J • d' 1 and 0.5-1.0 g*m- J »d'’, respectively. Wetlands with a composted organic 
substrate remove acidity from mine water at rates of 3-9 g*m* 2, d*’. A model for the design and sizing of passive treatment 
systems is presented in this report. 


14. SUBJECT TERMS 

Acid mine drainage, constructed wetlands, anoxic limestone drains, iron, sulfate, manganese, water 
treatment 

IS. NUMBER OF PACES 

38 

16. PRICE COPE 

T7. SECURJT Y 

CLASSIFICATION OF REPORT 

lb. SECURITY CLASSIFICATION OF THIS 

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19. SECURITY CLASSIFICATION OF 

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ABSTRACT 


*7ir 


96 







































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97 










































CONTENTS 


Page 

Abstract. 

Introduction. 

Treatment of mine water . 2 

Background of passive treatment. 

Acknowledgments. 3 

Chapter 1. Materials and methods. 4 

Collection of water samples . 4 

Analysis of water samples. 4 

Analytical quality control . 4 

Flow rate measurements. 4 

Analysis of surface deposits . 4 

Chapter 2. Chemical and biological processes in passive treatment systems . 5 

Acidity. 5 

Alkalinity. 6 

Metal removal processes. 

Metal removal in aerobic environments. 7 

Iron oxidation and hydrolysis . 

Manganese oxidation and hydrolysis . 8 

Mine water chemistry in anaerobic environments. 10 

Limestone dissolution. 10 

Sulfate reduction . 11 

Aluminum reactions in mine water. 13 

Chapter 3. Removal of contaminants by passive treatment systems . 14 

Evaluation of treatment system performance. 14 

Dilution adjustments . 16 

Loading limitations. 17 

Study sites. 18 

Effects of treatment systems on contaminant concentrations. 19 

Dilution factors . 19 

Removal of metals from alkaline mine water. 21 

Removal of metals and acidity from acid mine drainage. 23 

Chapter 4. Design and sizing of passive treatment systems . 25 

Characterization of mine drainage discharges . 26 

Calculations of contaminant loadings. 27 

Classification of discharges. 27 

Passive treatment of net alkaline water. 27 

Passive treatment of net acid water. 28 

Pretreatment of acidic water with ALD. 28 

Treating mine water with compost wetland. 30 

Operation and maintenance . 31 

Chapter 5. Summary and conclusions. 31 

Kinetics of contaminant removal processes . 32 

Long-term performance . 32 

Continually evolving passive technologies. 33 

References. 34 

ILLUSTRATIONS 

1. Comparison of calculated and measured acidities for water samples collected at Friendship Hill wetland 6 

2. Removal of Fe 2 * from acidic and alkaline mine waters in a laboratory experiment . 8 

3. Concentrations of Fe 3+ and field pH for water samples collected from Emlenton wetland. 8 

















































ii 

ILLUSTRATIONS—Continued 

Page 

4. Concentrations of Fe lc * and field pH at two constructed wetlands... 9 

5. Mean concentrations of Fe, Mn, and Mg at the Morrison wetland. 9 

6. Changes in concentrations of Fe 2 * and Mn 2 *. 10 

7. Concentrations of Ca, Fe, and alkalinity for water as it flows through the Howe Bridge ALD. 11 

8. Influent and effluent concentations at the Latrobe wetland . 14 

9. Relationship between mean Fe removal rates and .4, mean influent pH and B, mean influent alkalinity .. 22 

10. Relationship between mean Mn removal rates and A, mean influent pH and B , mean influent alkalinity . 22 

11. Measured rates of alkalinity generation and acidity removal at the Friendship Hill wetland. 25 

12. Flow chart showing chemical determinations necessary for the design of passive treatment systems. 26 

13. Longitudinal-section and cross-section of the Morrison ALD. 29 

TABLES 

1. Federal effluent limitations for coal mine drainage. 2 

2. Calculated and measured acidities for synthetic acidic mine water. 6 

3. Acidic components of mine drainage influent at three passive treatment systems . 6 

4. Chemical compositions of mine drainages that contain high concentrations of alkalinity. 7 

5. Chemistry of mine water flowing through the Howe Bridge anoxic limestone drain, January 23, 1992 ... 11 

6. Solubility products of some metal sulfides. 12 

7. Sinks for HjS in constructed wetlands and their net effect on mine water acidity and alkalinity. 12 

8. Surface and pore water chemistry at the Latrobe Wetland. 13 

9. Hypothetical wetland data and performance evaluations. 15 

10. Influent and effluent concentrations of Ca, Mg, Na, and sulfate at eight constructed wetlands. 16 

11. Average concentrations of Fe, Mn, and Mg at the Morrison passive treatment system . 17 

12. Construction characteristics of the constructed wetlands . 18 

13. Average chemical characteristics of influent water at the constructed wetlands. 19 

14. Mean water quality for sampling stations at the constructed wetlands . 20 

15. Dilution factors for the constructed wetlands . 21 

16. Fe and Mn removal rates at the constructed wetland. 21 

17. Fe and S0 4 content of ferric oxyhydroxide deposits. 24 

18. Average rates of acidity removal, sulfate removal, and calcium addition at sites receiving acidic mine 

water . 24 

19. Recommended sizing for passive treatment systems . 25 






























UNIT OF MEASURE ABBREVIATIONS USED IN THIS REPORT 


cm 

centimeter 

L^min* 1 

liter per minute 

°C 

degree Celsius 

m 

meter 

ft 

foot 

m 2 

square meter 

g 

gram 

fim 

micrometer 

g»cm' 3 

gram per cubic centimeter 

meq 

milliequivalent 

g*d‘ 

gram per day 

mg 

milligram 

g.m* 2 

gram per square meter 

mg»L _1 

milligram per liter 

g^m'^d' 1 

gram per square meter per day 

mg*L _1 »h' 1 

milligram per liter per hour 

g»m' 2 »yr -1 

gram per square meter per year 

mL 

milliliter 

gpm 

gallon per minute 

min 

minute 

ha 

hectare 

nmol 

nanomole 

h 

kg 

kg*d _1 

kg»m' 3 

hour 

kilogram (concentration) 

kilogram per day 

kilogram per cubic meter 

nmol»cm‘ 3 «d' 1 

yd 2 

yr 

nanomole per cubic centimeter 
per day 

square yard 

year 


L liter 




PASSIVE TREATMENT OF COAL MINE DRAINAGE 

By Robert S. Hedin , 1 Robert W. Nairn , 2 and Robert L P. Kleinmann 3 


ABSTRACT 

Passive methods of treating mine water use chemical and biological processes that decrease metal 
concentrations and neutralize acidity. Compared with conventional chemical treatment, passive methods 
generally require more land area, but use less costly reagents and require less operational attention and 
maintenance. Currently, three types of passive technologies exist: aerobic wetlands, organic substrate 
wetlands, and anoxic limestone drains. Aerobic wetlands promote mixed oxidation and hydrolysis 
reactions, and are most effective when the raw mine water is net alkaline. Organic substrate wetlands 
promote anaerobic bacterial activity that results in the precipitation of metal sulfides and the generation 
of bicarbonate alkalinity. Anoxic limestone drains generate bicarbonate alkalinity and can be useful for 
the pretreatment of mine water before it flows into a wetland. 

Rates of metal and acidity removal for passive systems have been developed empirically by the U.S. 
Bureau of Mines. Aerobic wetlands remove Fe and Mn from alkaline water at rates of 10-20 and 0.5- 
1.0 g»m' J «d' 1 , respectively. Wetlands with a composted organic substrate remove acidity from mine 
water at rates of 3-9 g«m‘ 2 *d" 1 . A model for the design and sizing of passive treatment systems is 
presented in this report. 


1 Rcscarch biologist. 

2 Re search biologist (now with The Ohio State University, Columbus, OH). 

3 Research supervisor. 

Pittsburgh Research Center, U.S. Bureau of Mines, Pittsburgh, PA. 





2 


INTRODUCTION 


TREATMENT OF MINE WATER 

The mining of coal in the Eastern and Midwestern 
United States can result in drainage that is contaminated 
with high concentrations of dissolved iron, manganese, 
aluminum, and sulfate. At sites mined since May 4,1984, 
drainage chemistry must meet strict effluent quality criteria 
(table 1). To meet these criteria, mining companies com¬ 
monly treat contaminated drainage using chemical meth¬ 
ods. In most treatment systems, metal contaminants are 
removed through the addition of alkaline chemicals (e.g., 
sodium hydroxide, calcium hydroxide, calcium oxide, sodi¬ 
um carbonate or ammonia). The chemicals used in these 
treatment systems can be expensive, especially when re¬ 
quired in large quantities. In addition, there are operation 
and maintenance costs associated with aeration and mixing 
devices, and additional costs associated with the disposal 
of metal-laden sludges that accumulate in settling ponds. 
It is not unusual for the water treatment costs to exceed 
$10,000 per year at sites that are otherwise successfully 
reclaimed. Total water treatment costs for the coal mining 
industry are estimated to exceed $1,000,000 per day (l). 4 
The high costs of water treatment place a serious financial 
burden on active mining companies and have contributed 
to the bankruptcies of many others. 


Table 1.—Federal effluent limitation* for coal mine drainage 


Pollutant or 

Maximum tor any 

Average of daily values 

pollutant 

1 day, 

for 30 consecutive 

property 

mg-L' 1 

days mg-l' 1 

Fe total . 

6.0 

3.0 

Mn total. 

4.0 

2.0 


pH between 6.0 and 9.0. 


The high costs of chemical systems also limit the water 
treatment efforts at abandoned sites. Thousands of miles 
of streams and rivers in Appalachia are currently polluted 
by the input of mine drainage from sites that were mined 
and abandoned before enactment of strict effluent regula¬ 
tions (2-3). State and Federal reclamation agencies, local 
conservation organizations, and watershed associations all 
consider the treatment of contaminated coal mine dis¬ 
charges to be a high priority. Unfortunately, insufficient 
funds are available for chemical water treatment, except in 
a few watersheds of special value. 

Natural processes commonly ameliorate mine drainage 
pollution. As contaminated coal mine drainage flows into 
and through receiving systems (streams, rivers, and lakes), 

4 Italic numbers in parentheses refer to items in the list of references 
at the end of this report 


its toxic characteristics decrease naturally as a result of 
chemical and biological reactions and by dilution with 
uncontaminated water. The low pH that is common to 
many mine drainages is raised when the water mixes with 
less acidic or alkaline water or through direct contact with 
carbonate rocks. Metal contaminants of coal mine 
drainage then precipitate as oxides and hydroxides under 
the aerobic conditions found in most surface waters. Dis¬ 
solved Fe precipitates as an oxyhydroxide, staining the 
bottoms of many streams orange and often accumulating 
to sufficient depths to suffocate benthic organisms. Less 
commonly, dissolved Mn precipitates as an oxide that 
stains rocks and detrital material black. Dissolved A1 
precipitates as a white hydroxide. 

During the last decade, the possibility that mine water 
might be treated passively has developed from an experi¬ 
mental concept to full-scale field implementation at hun¬ 
dreds of sites. Passive technologies take advantage of 
natural chemical and biological processes that ameliorate 
contaminated water conditions. Ideally, passive treatment 
systems require no input of chemicals and little or no 
operation and maintenance requirements. The costs of 
passive treatment systems are generally measured in their 
land use requirements. Passive treatment systems use con¬ 
taminant removal processes that are slower than that of 
conventional treatment and thus require longer retention 
times and larger areas to achieve similar results. 

The goal of passive mine drainage treatment systems 
is to enhance the natural amelioration processes so that 
they occur within the treatment system, not in the re¬ 
ceiving water body. Two factors that determine whether 
this goal can be accomplished are the kinetics of the 
contaminant removal processes and the retention time of 
the mine water in the treatment system. The retention 
time for a particular minesite is often limited by available 
land area. However, the kinetics of contaminant removal 
processes can often be affected by manipulating the 
environmental conditions that exist within the passive 
treatment system. Efficient manipulation of contaminant 
removal processes requires that the nature of the rate- 
limiting aspects of each removal process be understood. 

This U.S. Bureau of Mines (USBM) report describes 
the chemical and biological processes that underlie the 
passive technologies currently used in the eastern United 
States for the treatment of contaminated coal mine 
drainage. After reviewing the background of passive treat¬ 
ment and the methods used in these studies (Chapter 1), 
the chemical behavior of mine drainage contaminants is 
reviewed (Chapter 2). This discussion highlights the dif¬ 
ference between alkaline and acidic mine water, and de¬ 
tails the processes in passive treatment systems that 
generate alkalinity. In Chapter 3, contaminant removal is 









3 


evaluated for 13 passive treatment systems through the 
calculation of contaminant removal rates. These rates, 
which incorporate the size of the treatment system, the 
flow rate of the water, and mine drainage chemistry, are 
the only measures of treatment system performance that 
can be reliably compared between systems. In Chapter 4, 
the chemical background provided in Chapter 2 and the 
observed contaminant removal rates presented in Chap¬ 
ter 3 are combined in a model that gives design and sizing 
recommendations for future passive treatment systems. 
Chapter 5 summarizes the results of this study and iden¬ 
tifies future research needs. 

BACKGROUND OF PASSIVE TREATMENT 

The current interest in passive treatment technologies 
can be traced to two independent research projects that 
indicated that natural Sphagnum wetlands caused an 
amelioration of mine drainage pollution without incurring 
any obvious ecological damage (4-5). These observations 
prompted the idea that wetlands might be constructed for 
the intentional treatment of coal mine drainage. Research 
efforts were initiated by West Virginia University, Wright 
State University, Pennsylvania State University, and the 
USBM to evaluate the feasibility of the idea. As a result 
of promising preliminary reports (6-8), experimental wet¬ 
lands were built by mining companies and reclamation 
groups. Initially, most of these wetlands were constructed 
to mimic Sphagnum moss wetlands. However, Sphagnum 
moss was not readily available, proved difficult to trans¬ 
plant, and tended to accumulate metals to levels that were 
toxic to the Sphagnum after several months of exposure to 
mine drainage (9-10). Instead of abandoning the concept, 
researchers experimented with different kinds of con¬ 
structed wetlands. Eventually, a wetland design evolved 
that proved tolerant to years of exposure to contaminated 
mine drainage and was effective at lowering concentrations 
of dissolved metals. Most of these treatment systems con¬ 
sist of a series of small wetlands (< 1 ha) that are vege¬ 
tated with cattails (Typha latifolia) (11-12). In northern 
Appalachia, many wetlands contain a compost and lime¬ 
stone substrate in which the cattails root. In southern 
Appalachia, most wetlands have been constructed without 
an exogenous organic substrate; emergent plants have been 
rooted in whatever soil or spoil substrate was available on 
the site when the treatment system was constructed (13). 


Recently, treatment technologies have been developed 
that do not rely at all on the wetland model that the early 
systems were designed to mimic. Ponds, ditches, and rock- 
filled basins have been constructed that are not planted 
with emergent plarits, and in some cases, contain no soil or 
organic substrate (14). Pretreatment systems have been 
developed where acidic water contacts limestone in an 
anoxic environment before flowing into a settling pond or 
wetland system (15). In these cases, the water is treated 
with limestone followed by passive aeration; however, the 
low cost and chemical behavior of limestone make possible 
the construction of wetland systems that should, theo¬ 
retically, require no maintenance and last for decades. 

A wide diversity of opinions exist on the merits of pas¬ 
sive treatment systems for mine drainage. Wieder’s anal¬ 
ysis of a survey of constructed wetlands conducted by the 
Office of Surface Mining (OSM) indicated no strong re¬ 
lationships between concentration efficiency and wetland 
design features, leading him to question the feasibility of 
the constructed wetland concept (12). In a separate study 
by Wieder and his colleagues, measurements of the Fe 
content of Sphagnum peat exposed to synthetic acid mine 
drainage were used to calculate that an average wetland 
system should cease to remove metals after 11 weeks of 
operation (16). These negative reports contrast with many 
other studies of successful wetlands. Examples include an 
Ohio wetland that is treating Fe-contaminated mine 
drainage effectively in its 8th year of operation (17) and six 
Tennessee Valley Authority (TVA) wetlands that have 
produced compliance water for at least 4 years (18). A 
vast majority of the passive treatment systems constructed 
in the United States during the last decade achieve per¬ 
formance that is better than Wieder and his colleagues 
would predict, though not necessarily enough to consist¬ 
ently meet effluent limits. Hundreds of constructed wet¬ 
lands discharge water that contains lower concentrations 
of metal contaminants than was contained in the inflow 
drainage. These improvements in water quality decrease 
the costs of subsequent water treatment at active sites and 
decrease deleterious impacts that discharges from aban¬ 
doned sites have on receiving streams and lakes. In gen¬ 
eral, the systems that are not 100% effective were im¬ 
properly designed, were undersized, or both. This report 
has been prepared so that designers of future systems can 
avoid these errors. 


ACKNOWLEDGMENTS 


The authors thank Holly Biddle for invaluable assist¬ 
ance with a reorganization of this report that occurred be¬ 
tween draft and final versions. Laboratory analyses were 
conducted by Mark Wesolowski, Joyce Swank, and Dennis 


Viscusi. Adrian Woods, John Odoski, John Kleinhenz, 
and Robert Neupert assisted with field work. Partial 
funding for research described in this report was provided 
by the U.S. Office of Surface Mining. 


4 


CHAPTER 1. MATERIALS AND METHODS 


COLLECTION OF WATER SAMPLES 

Water samples were collected at passive treatment 
systems from their influent and effluent points, and, if 
applicable, between treatment cells within the system. 
Raw and acidified (2 mL of concentrated HCl) water sam¬ 
ples were collected in 250 mL plastic bottles at each sam¬ 
pling point. Measurements of pH and temperature were 
made in the field with a calibrated Orion SA 270, SA 250 
or SA 290 portable pH/ISE meter. 5 Alkalinity was meas¬ 
ured in the field using a pH meter and an Orion Total 
Alkalinity Test Kit. At sites where particulates were vis¬ 
ible in water samples, an extra sample was collected that 
was filtered through a 0.22 -pun membrane filter before 
acidification. All samples were immediately placed on 
ice in an insulated cooler and returned to the laboratory 
within 36 h of collection. Samples were refrigerated at 
4° C until analysis. 

Substrate pore water samples were collected using a 
dialysis method similar to that described by Wheeler and 
Ciller (19). Lengths of 6,000-8,000 molecular weight 
dialysis tubing were filled with 250 mL of deionized, de- 
oxygenated water and buried 30-45 cm deep in the organic 
substrate of the wetland. Three weeks later, the dialysis 
tubes were retrieved and the contents immediately filtered 
through a 0.45-/um membrane filter. Laboratory experi¬ 
ments established that the chemistry of water within the 
sampling tubes equilibrated with surrounding pore water 
within 24 h. The 3-week equilibration period was allowed 
so that chemical anomalies caused by the burial process 
would dissipate. Portions of the filtered water samples 
were preserved with NaOH (for dissolved sulfide deter¬ 
minations), HCl (for cation analysis), or were left unpre¬ 
served (for alkalinity, acidity, and sulfate analyses). 

ANALYSIS OF WATER SAMPLES 

Concentrations of Fe, Mn, Al, Ca, Mg, and Na were 
determined in the acidified samples using Inductively 
Coupled Argon Plasma Spectroscopy, ICP (Instrumenta¬ 
tion Laboratory Plasma 100 model). The acidified samples 
were first filtered through a 0.45-^m membrane filter to 
prevent clogging of the small diameter tubing in the ICP. 

Ferrous iron concentrations were determined on acid¬ 
ified samples by the potassium dichromate method (20). 
Sulfate concentrations were determined by reaction with 

Reference lo specific products does not imply endorsement by the 
U.S. Bureau of Mines. 


barium chloride (BaCl) after first passing the raw sample 
through a cation exchange resin. Thorin was used as the 
end-point indicator. Dissolved sulfide species were deter¬ 
mined using a sulfide-specific electrode. 

Acidity was determined by boiling a 50-mL raw sample 
with 1 mL of 30% H 2 0 2 (hydrogen peroxide), and then 
titrating the solution with 0.1 N NaOH (sodium hydroxide) 
to pH 8.3 (27). Acidity and alkalinity are reported as 
mg»L _1 CaC0 3 equivalents. 

ANALYTICAL QUALITY CONTROL 

For each set of samples for a particular site, a dupli¬ 
cate, standard, and spike were analyzed for quality control 
purposes. The relative standard deviation for the duplicate 
was always at least 95%. Percent recovery for the stand¬ 
ards were within 3% of the original standard. Spike recov¬ 
eries were within 5% of the expected values. 

FLOW RATE MEASUREMENTS 

Mine water flow rates were determined by several 
methods. Whenever possible, flow was determined with a 
bucket and stopwatch. In all cases, three to five meas¬ 
urements of the time needed to collect a known volume 
of water were made at each sampling location, and the 
average flow rate of these measurements was reported. At 
two sites where flows were occasionally too high to meas¬ 
ure with a bucket (the Latrobe and Piney Wetlands), 0.50 
or 0.75 ft H-typc flumes were installed and flows were 
determined from the depth of water in the flume. At the 
Keystone site, flows were determined by measuring the 
depth of water in a drainage pipe and then using the 
Manning formula for measurement of gravity flow in open 
channels (22). 

ANALYSIS OF SURFACE DEPOSITS 

The chemical composition of surface deposits collected 
from several constructed wetlands were determined by the 
following procedure. The samples were rinsed with 
deionized water, dried at 100° C, and weighed. The acid- 
soluble component was extracted by boiling 5 g of dry 
sample in 20 mL of concentrated HCl for 2 min. The acid 
extractants were filtered and analyzed for metal content 
by ICP Spectroscopy and for sulfate content by liquid 
chromatography. The acid-insoluble material was dried at 
100* C and weighed. The acid-soluble component was 
determined by subtracting the dry weight of the insoluble 
material from the original dry weight. 



CHAPTER 2. CHEMICAL AND BIOLOGICAL PROCESSES 
IN PASSIVE TREATMENT SYSTEMS 


5 


Coal mining can promote pyrite oxidation and result 
in drainage containing high concentrations of Fe, Mn, and 
Al, as well as S0 4 , Ca, Mg, and Na. The solubilities of Fe, 
Mn, and Al are generally very low (<1 mg»L' 1 ) in nat¬ 
ural waters because of chemical and biological processes 
that cause their precipitation in surface water environ¬ 
ments. The same chemical and biological processes re¬ 
move Fe, Mn, and Al from contaminated coal mine drain¬ 
age, but the metal loadings from abandoned minesites are 
often so high that the deleterious effects of these elements 
persist long enough to result in the pollution of receiving 
waters. 

Passive treatment systems function by retaining con¬ 
taminated mine water long enough to decrease contam¬ 
inant concentrations to acceptable levels. The chemical 
and biological processes that remove contaminants vary 
between metals and are affected by the mine water pH 
and oxidation-reduction potential (Eh). Efficient passive 
treatment systems create conditions that promote the 
processes that most rapidly remove target contaminants. 
Thus, the design of passive treatment systems must be 
based on a solid understanding of mine drainage chem¬ 
istry and how different passive technologies affect this 
chemistry. 

This chapter provides the basic chemical and biological 
background necessary to efficiently design passive treat¬ 
ment systems. The authors begin with a discussion of 
acidity and alkalinity because many of the decisions about 
how to treat mine water passively depend on determina¬ 
tions of these parameters. Next, the chemistry of Fe, Mn, 
and Al in aerobic and anaerobic aquatic environments is 
described. Throughout the discussion, chemical and bio¬ 
logical concepts are illustrated with data collected from 
passive treatment systems. 

ACIDITY 

Acidity is a measurement of the base neutralization 
capacity of a volume of water. Three types of acidity exist: 
proton acidity associated with pH (a measure of free H* 
ions), organic acidity associated with dissolved organic 
compounds, and mineral acidity associated with dissolved 
metals (23). Mine waters generally have a very low dis¬ 
solved organic carbon content, so organic acidity is very 
low. The acidity of coal mine drainage arises from free 
protons (low pH) and the mineral acidity from dissolved 
Fe, Mn, and Al. These metals are considered acidic be¬ 
cause they can undergo hydrolysis reactions that produce 
H\ 


Fe r + 1/40 2 + 3/2H 2 0 - FeOOH + 2H* (A) 
Fe 3 * + 2H 2 0 - FeOOH + 3H* (B) 

Al 3 * + 3H 2 0 - Al(OH) 3 + 3H* (C) 

Mn 2 * + 1/40 2 ♦ 3/2H 2 0 - MnOOH ♦ 2H* (D) 

These reactions can be used to calculate the total 
acidity of a mine water sample and to partition the acidity 
into its various components. The expected acidity of a 
mine water sample is calculated from its pH and the sum 
of the milliequivalents of acidic metals. For most coal 
mine drainages, the calculation is as follows: 

Acid^ = 50(2Fe 2+ /56 ♦ 3Fe 3 756 (1) 

♦ 3A1/27 + 2Mn/55 + 1000(10"P H )) 

where all metal concentrations are in milligram per liter 
and 50 is the equivalent weight of CaC0 3 , and thus trans¬ 
forms milliequivalent per liter of acidity into milligram per 
liter CaC0 3 equivalent. For water samples with pH <4.5 
(no alkalinity present), equation 1 cakulates a mine water 
acidity that corresponds closely with measurements of 
acidity made using the standard H 2 0 2 method (21). Using 
synthetic mine drainages with a wide range of composi¬ 
tions, it was determined that calculated acidities differed 
from measured values by less than 10% (table 2). 

Equation 1 accurately characterizes mineral acidity for 
samples of actual add mine drainage as well. At one site 
where numerous measurements of metal chemistry and 
total acidity were made, the mean acidity of samples with 
pH <4.5 was 693 mg'L' 1 , while the predicted acidities for 
these samples averaged 655 mg^L* 1 , a difference of only 
6% (figure 1). 

Equation 1 can be used to partition total acidity into its 
individual constituents. When the total acidities of con¬ 
taminated coal mine drainages are partitioned in this 
manner, the importance of mineral aridity becomes ap¬ 
parent. A breakdown of the acidic components of three 
mine drainages is shown in table 3. At each site, the acid¬ 
ity arising from protons (pH) was a minor component of 
the total aridity. Mine drainage at the Friendship Hill 
wetland had extremely low pH (2.7), but the acidity of the 


6 



Figure 1.—Comparison of calculated and measured acidities 
for water samples collected at Friendship Hill wetland. 


mine water resulted primarily from dissolved ferric iron 
and Al. The Somerset wetland received water with low 
pH (3.7), but the acidity of the water resulted largely from 
dissolved ferrous iron and Mn. At the Cedar Grove sys¬ 
tem, where the mine water was circumneutral, ferrous iron 
accounted for 98% of the acidity, while the hydrogen ion 
accounted for < 1% of mine water acidity. 


ALKALINITY 

When mine water has pH >4.5, it has acid neutralizing 
capacity and is said to contain alkalinity. Alkalinity can 
result from hydroxyl ion (OH'), carbonate, silicate, bo¬ 
rate, organic ligands, phosphate, and ammonia (23). The 
principal source of alkalinity in mine water is dissolved 
carbonate, which can exist in a bicarbonate (HC0 3 ') or 
carbonate form (C0 3 2 ). Both can neutralize proton 
acidity. 

H + + HC0 3 _ - H 2 0 + C0 2 (E) 

2H* + C0 3 ' -* H 2 0 + C0 2 (F) 

In the pH range of most alkaline mine waters (5 to 8), 
bicarbonate is the principal source of alkalinity. 

The presence of bicarbonate alkalinity in mine waters 
that contain elevated levels of metals is not unusual. 
Table 4 shows the chemical composition of 12 mine waters 
in northern Appalachia that contain alkalinity and are also 
contaminated with ferrous iron and Mn. None are con¬ 
taminated with dissolved ferric iron or Al because the 
solubility of these metals is low in mine waters with pH 
greater than 5.5 (23-24). 


Table 2.—Calculated and measured acidities for synthetic acidic mine water 


Synthetic Mine Water Composition 1 


Acidity 


pH 

Fe 2 * 

Fe 3 * 

Al 

Mn 

Calculated 2 

Measured 3 

Diff. 4 

3.9 

98 

1 

0 

0 

181 

184 

-2% 

3.9 

0 

0 

106 

0 

598 

578 

+ 3% 

3.6 

0 

0 

0 

97 

192 

186 

+ 3% 

3.8 

13 

0 

47 

42 

370 

335 

+ 9% 


Measured values are the average of three tests. Metal concentrations are 
mg-L' 1 . Acidities are mg'l 1 CaC0 3 equivalent. 

2 From reaction 1. 

3 0ata determined by the hot H 2 0 2 acidity method (27). 

4 (1.00 - meas/cal) x 100. 

Table 3.—Acidic components of mine drainage influent at three passive treatment systems 


Parameter 


Friendship Hill 



Somerset 



Cedar Grove 


Concen¬ 

tration, 

mg-L' 1 

Acid 

equivalent, 1 

mg-L' 1 

% of 
total 
acidity 

Concen¬ 

tration, 

mg-L' 1 

Acid 

equivalent, 1 

mg-L' 1 

% of 
total 
acidity 

Concen¬ 

tration, 

mg-L' 1 

Acid 

equivalent, 1 

mg-L' 1 

% of 
total 
acidity 

Fe 2+ . 

7 

13 

1 

193 

345 

69 

95 

170 

98 

Fe 3+ . 

153 

434 

49 

9 

24 

5 

<1 

<1 

<1 

Al 3+ . 

58 

317 

36 

3 

17 

3 

<1 

<1 

<1 

Mn 2+ . 

9 

16 

1 

59 

107 

21 

2 

4 

2 

PH. 

2.6 

112 

13 

3.7 

10 

2 

6.3 

<1 

<1 


'CaC0 3 equivalents calculated from the stoichiometry of reactions A-D. 























7 


Table 4.—Chemical compositions of mine drainages that contain high concentrations of alkalinity 


Location 

pH 

Alkalinity, 

mg-L 1 

Al, 

mg-L 1 

Fe 2 *, 

mg-L 4 

Fe 3 *, 

mg-L 1 

Mn, 

mg-L' 1 

S0 4 , 

mg-L' 1 

Net alkalinity, 1 
mg-L' 1 

Ohio: Coshocton. 

Pennsylvania: 


152 

<1 

119 

<1 

2 

1,325 

-50 

Cross Creek. 


300 

<1 

96 

<1 

2 

1,260 

140 

Donegal. 


214 

<1 

39 

<1 

8 

830 

130 

Fallston. 

6.2 

120 

<1 

30 

<1 

3 

390 

66 

Keystone. 

6.5 

106 

<1 

37 

<1 

1 

331 

72 

Latrobe . 


204 

<1 

102 

<1 

6 

1,200 

15 

New Bethlehem. 

6.1 

163 

<1 

51 

<1 

28 

493 

51 

Possum Hollow. 

_ 6.4 

263 

<1 

32 

<1 

1 

620 

209 

Sligo. 


93 

<1 

43 

<1 

26 

1,720 

-31 

Somerset. 


275 

<1 

2 

<1 

6 

750 

265 

St. Petersburg. 


255 

<1 

29 

<1 

9 

250 

203 

Uniontown. 

6.3 

220 

<1 

70 

<1 

3 

950 

95 


'Alkalinity minus acidity. 


Alkalinity and acidity arc not mutually exclusive terms. 
All of the mine waters shown in table 4 contain both acid¬ 
ity and alkalinity. When water contains both mineral 
acidity and alkalinity, a comparison of the two measure¬ 
ments results in a determination as to whether the water 
is net alkaline (alkalinity greater than acidity) or net acidic 
(acidity greater than alkalinity). Net alkaline water con¬ 
tains enough alkalinity to neutralize the mineral acidity 
represented by dissolved ferrous iron and Mn. As these 
metals oxidize and hydrolyze, the proton acidity that is 
produced is rapidly neutralized by bicarbonate. For waters 
contaminated with Fe 2 *, the net reaction for the oxidation, 
hydrolysis and neutralization reactions is 


complex because it differs between metals and also 
between abiotic and biotic processes. 

METAL REMOVAL IN AEROBIC ENVIRONMENTS 
Iron Oxidation and Hydrolysis 

The most common contaminant of coal mine drainage 
is ferrous iron. In oxidizing environments common to 
most surface waters, ferrous iron is oxidized to ferric iron. 
Ferrous iron oxidation occurs both abiotically and as a 
result of bacterial activity. The stoichiometry of the reac¬ 
tion is the same for both oxidation processes. 


Fe 2 * + Y*0 2 + 2HC0 3 ' - FeOOH + Vi H 2 0 + 2C0 2 (G) 

Reaction G indicates that net alkaline waters contain 
at least 1.8 mg»L _1 alkalinity for each 1.0 mg»L"‘ of dis¬ 
solved Fe. Waters that contain a lesser ratio are net 
acidic, since the oxidation and hydrolysis of the total dis¬ 
solved iron content results in a net release of protons and 
a decrease in the pH. 

METAL REMOVAL PROCESSES 

Oxidation and hydrolysis reactions already discussed 
cause concentrations of Fe 2 *, Fe 3 *, Mn, and Al to com¬ 
monly decrease when mine water flows through an aerobic 
environment. Whether these reactions occur quickly 
enough to lower metal concentrations to an acceptable 
level depends on the availability of oxygen for oxidation 
reactions, the pH of the water, the activity of microbial 
catalysts, and the retention time of water in the treatment 
system. The pH is an especially important parameter 
because it influences both the solubility of metal hydrox¬ 
ide precipitates and the kinetics of the oxidation and 
hydrolysis processes. The relationship between pH and 
metal-removal processes in passive treatment systems is 


Fe 2 * + ‘/«0 2 + H* - Fe 3 * + 1/2H 2 0 (H) 

The pH of the mine water affects the kinetics of both the 
abiotic and biotic processes (25-26). When oxygen is not 
limiting, the rate of abiotic Fe oxidation slows 100-fold for 
every unit decrease in pH. At pH values >8, the abiotic 
process is fast (rates are measured in seconds), while at 
pH values <5 the abiotic process is slow (rates are 
measured in days). In contrast, bacterial oxidation of 
ferrous iron peaks at pH values between 2 and 3, while 
less activity occurs at pH values >5 (27). The presence of 
bicarbonate alkalinity buffers mine water at a pH of 6 to 
7, a range at which abiotic iron oxidation processes should 
dominate. Waters containing no alkalinity have a pH <4.5 
and the removal of Fe under oxidizing conditions occurs 
primarily by bacterial oxidation accompanied by hydrolysis 
and precipitation. 

The effect that pH can have on the mechanism of iron 
oxidation is shown by the data in figure 2. Samples were 
collected from two mine drainages that were both con¬ 
taminated with ferrous iron, but had different pH and 
alkalinity values. The samples were returned to the lab¬ 
oratory and exposed to aerobic conditions. For the cir- 
cumncutral waters, oxidation of ferrous iron occurred at a 


















8 




Figure 2.—Removal of Fe 2 ’ from acidic and alkaline mine 
water* In laboratory experiment Raw mine drainage was col¬ 
lected from A, acidic Latrobe site; 8, alkaline Cedar Grove site. 
Splits of each sample were filter-sterilized (0J22-nm Alter). The 
Latrobe sample* were shaken throughout experiment; air was 
bubbled through Cedar Grove samples during experiment 

rate of 18 mg»L' 1 «h' 1 , while the rate for the raw acidic 
samples was only 1.4 mg’L'^h' 1 . To evaluate the signi¬ 
ficance of bacterial processes in iron oxidation, splits of 
both samples were filter-sterilized ( 0.22-fim membrane 
filter) before the experiment was begun. Removal of bac¬ 
teria had no effect on the oxidation of ferrous iron for the 
circumneutral water, but completely inhibited ferrous iron 
oxidation for the acidic water. 

As ferrous iron is converted to ferric iron, it is sub¬ 
ject to hydrolysis reactions that can precipitate it as a 
hydroxide (reaction B). The hydrolysis reaction occurs 
abiotically, catalysis of the reaction by microorganisms has 
not been demonstrated. The solubility of the ferric hy¬ 
droxide solid is such that, under equilibrium conditions, 
negligible dissolved ferric iron (<1 mg’L' 1 ) exists unless 
the pH of the mine water is <2.5. In actuality, the rate of 
the hydrolysis reaction is also pH dependent, and sig¬ 
nificant Fe 3 * can be found in mine water with a pH above 
2.5. Singer and Stumm (25) suggested a fourth-order rela¬ 
tionship with pH, which indicated that ferric iron hydrol¬ 
ysis processes shift from a very rapid rate at pH >3 to a 
very slow rate at pH <2.5. Figure 3 shows the relation¬ 
ship between pH and concentrations of Fe 3 * at a site 
where pH varied by almost 3 units. Ferric iron was not 
generally indicated unless the pH was <4, and the highest 


40 

- 1 - 1 —— - 1 - r— - 

35 

• 

30 

• 

L 25 

a • 

| 20 

*• *if. i * 

4-* 15 


ro 


Fe 

o 

* Y* 

5 

<4 • 

• • • 



V 

- 1_1-1 ! ..1- 


FIELD, pH 

Figurs 3.— Concentrations of F# 3 * and field pH for water 
samples collected from Emlenton wetland. 

concentrations of ferric iron occurred when the pH was 
<3. 

The tendency for dissolved iron to oxidize and hydro¬ 
lyze in aerobic environments with pH >3 results in the 
precipitation of ferric hydroxide. Because the net result of 
the oxidation and hydrolysis process is the production of 
protons, the process can decrease pH. Thus, natural or 
constructed wetlands receiving circumneutral net acidic 
water commonly decrease both Fe concentrations and pH. 
An example of this phenomenon is shown in figure 44. 
As water flowed through the constructed wetland, iron 
concentrations decreased from 95 to 15 mg*L _1 , and pH 
decreased from 5.5 to 3.2. Figure 42? shows Fe concen¬ 
trations and pH within a wetland that received mine water 
with a net alkalinity. Despite the removal of 60 mg’L' 1 
Fe 2 * and the production of enough protons to theoret¬ 
ically lower the pH to 2.7, the pH did not decrease 
because bicarbonate alkalinity neutralized the proton 
acidity. 

Manganese Oxidation and Hydrolysis 

Manganese undergoes oxidation and hydrolysis reac¬ 
tions that result in the precipitation of manganese oxy- 
hydroxides. The specific mechanism(s) by which Mn 2 ' 
precipitates from aerobic mine water in the absence of 
chemical additions is uncertain. Mn 2 * may be oxidized to 
either a +3 or a +4 valance, either one of which rapidly 
precipitates (reaction D). If MnOOH precipitates, over 
time it likely oxidizes to the more stable MnO z . In alka¬ 
line environments, Mn 2 * can precipitate as a carbonate, 
which may also be oxidized by oxygen to Mn0 2 (25). 

Mn 2 * + HC0 3 - MnC0 3 + H + (I) 

MnC0 3 + V0 2 -* Mn0 2 + C0 2 (J) 









9 


Regardless of the mechanism by which Mn 2 * is oxidized 
to Mn 4 *, the removal of one mole of Mn 2 * from solution 
results in the release of two moles of H* or an equivalent 
decrease in alkalinity (HC0 3 *). 

The kinetics of Mn 2 * oxidation reactions are strongly 
affected by pH. Abiotic oxidation reactions are very slow 
at pH <8 (24). Microorganisms can catalyze Mn 2 * oxida¬ 
tion, but their activity is limited to aerobic waters with pH 
>6 (29). ' 

Although the hydrolysis of Mn produces protons, the 
precipitation of MnOOH does not result in large declines 
in pH as can happen when FeOOH precipitates. This dif¬ 
ference between Mn and Fe chemistry is because of the 
fact that no natural mechanism exists that rapidly oxidizes 
Mn 2 * under acidic conditions. If pH falls below 6, Mn 2 * 
oxidation virtually ceases, the proton-producing hydrolysis 
reaction ceases, and pH stabilizes. 

The oxidation and precipitation of Mn 2 * from solution 
is accelerated by the presence of Mn0 2 and FeOOH (24, 
30). Both solids reportedly act as adsorption surfaces for 
Mn 2 * and catalyze the oxidation mechanism. While addi¬ 
tions of FeOOH to Mn-containing water might accelerate 
Mn oxidation, the direct precipitation of FeOOH from 
mine water containing Fe 2 * does not generally stimulate 



SAMPLING STATION 

Figure 4.—Concentration* of Fe' 0 * end field pH at two con¬ 
structed wetlands. A, Emlenton wetland; B, Cedar Grove wetland. 


Mn-removal processes in passive treatment systems. Fig¬ 
ure 5 shows concentrations of Mn and Fe for mine water 
as it flowed through a constructed wetland that markedly 
decreased concentrations of both metals. On average, Fe 
decreased from 150 to <1 mg*L~ l , while Mn decreased 
from 42 to 11 mg»L* 1 . Removal of metals occurred se¬ 
quentially, not simultaneously. Two-thirds of the decrease 
in iron concentration occurred between the first and 
second sampling stations. The wetland substrate in this 
area was covered with precipitated FeOOH and the water 
was turbid with suspended FeOOH. Despite the presence 
of large quantities of FeOOH, little change in the con¬ 
centration of Mn occurred between the first and second 
sampling station. The slight decrease in Mn that occurred 
was proportionally similar to the change in Mg, suggesting 
that dilution was the most likely cause of the decrease in 
Mn concentrations (the use of Mg to estimate dilution is 
discussed in detail in chapter 3). Between stations 3 and 
5, there was little Fe present in the water and little visual 
evidence of FeOOH sludge on the wetland substrate. 
Most of the observed removal of Mn occurred in this Fe- 
free zone. 

The absence of simultaneous precipitation of dissolved 
Fe and Mn from aerobic alkaline waters likely results from 
the reduction of oxidized forms of Mn by ferrous iron. 

Mn0 2 + 2Fe 2 * + 2H z O - 2FeOOH + Mn 2 * + 2H* (K) 
or 

MnOOH + Fe 2 * ^ FeOOH + Mn 2 * (L) 

Figure 6 shows the results of a laboratory study that 
demonstrate the instability of Mn oxides in the presence 
of ferrous iron. Water samples and Mn-oxides were 



Figure 5.—Mean concentrations of Fe, Mn, and Mg at the 
Morrison Wetland. Mine water flows linearly from station 1 to 
station 5. Verticte bars are one standard error of the mean. 






10 



0 20 40 60 80 100 

TIME, h 


Figure 6.—Changes in concentrations of Fe 2 * and Mn 2 *. A, 
absence; B, presence of MnOOH. Mine water was collected from 
influent pipe of Biair wetland. MnOOH was collected from inside 
of final effluent pipe. 

collected from a wetland that removed Fe and Mn in a 
sequential manner. The wetland influent was alkaline 
(pH 6.2, 162 mg'L' 1 alkalinity) and contaminated with 
50 mg'L* 1 Fe and 32 mg^L* 1 Mn. Two flasks of mine 
water received MnO z additions, while the controls did not 
receive Mn0 2 . Concentrations of dissolved Fe and Mn 
were monitored in each flask over a 73-h period. In all 
flasks, concentrations of Fe decreased to <1 mg»L _1 . In 
the control flasks, concentrations of Fe decreased to 
<3 mg’L' 1 within 43 h. In flasks that received Mn0 2 , 
concentrations of Fe decreased to <3 mg’L' 1 in only 
22 h. No change in concentrations of Mn occurred in the 
control flasks. Concentrations of Mn in the Mn0 2 flasks 
increased by 15 mg*L _1 during the first 22 h and did not 
change during the remaining 50 h of the experiment. The 
association of accelerated precipitation of Fe with 
solubilization of Mn 2 * suggests that the Mn0 2 oxidized 
Fe 2 * in a manner analogous to reaction K. 

The data presented in figures 5 and 6 demonstrate 
aspects of Fe and Mn chemistry that are important in 
passive treatment systems. Iron oxidizes and precipitates 
from alkaline mine water much more rapidly than does 
Mn. One reason for the differences in kinetics is that the 


oxidized Mn solids, which are presumed to result from 
Mn 2 * oxidation reactions, are not stable in the presence 
of Fe 2 *. Concentrations of ferrous iron must decrease to 
very low levels before Mn 2 * oxidation processes can result 
in a stable solid precipitate. In the absence of Fe 2 *, Mn 
removal is still a very slow process under laboratory con¬ 
ditions. Conditions in a wetland may either accelerate 
Mn-removal reactions or promote mechanisms that are not 
simulated in simple laboratory experiments. However, 
both field and laboratory investigations indicate that, under 
aerobic conditions, the removal of Mn occurs at a much 
slower rate than does the removal of Fe (empirical evi¬ 
dence for this concept is presented in chapter 3). 

MINE WATER CHEMISTRY IN ANAEROBIC 
ENVIRONMENTS 

Chemical and microbial processes in anaerobic envi¬ 
ronments differ from those observed in aerobic envi¬ 
ronments. Because 0 2 is absent, Fe 2 * and Mn 2 * do not 
oxidize and oxyhydroxide precipitates do not form. Hy¬ 
droxides of the reduced Fe and Mn ions, Fe(OH) 2 and 
Mn(OH) 2> do not form because of their high solubility 
under acidic or circumneutral conditions. In passive treat¬ 
ment systems where mine water flows through anaerobic 
environments, its chemistry is affected by chemical and 
biological processes that generate bicarbonate and hydro¬ 
gen sulfide. 

Limestone Dissolution 

A major source of bicarbonate in many anaerobic en¬ 
vironments is the dissolution of carbonate minerals, such 
as calcite. 

CaC0 3 + H + - Ca 2+ -*• HC0 3 (M) 

Carbonate dissolution can result in higher concen¬ 
trations of bicarbonate in anaerobic mine water environ¬ 
ments than aerobic environments for two reasons. First, 
the absence of Fe 3 * in most anaerobic environments limits 
the formation of FeOOH coatings that armor carbonate 
surfaces and inhibit further carbonate dissolution in aero¬ 
bic environments (31). Second, the solubilities of carbon¬ 
ate compounds are directly affected by the partial pressure 
of dissolved C0 2 (23-24, 32). Anaerobic mine water en¬ 
vironments commonly contain high C0 2 partial pressures 
because of the decomposition of organic matter and the 
neutralization of proton acidity. 

The observation that limestone dissolution is enhanced 
when contact with mine water occurs in an anaerobic 
environment has resulted in the construction of anaerobic 
limestone treatment systems. The first demonstration of 
















11 


this technology was by Turner and McCoy (75) who 
showed that when anoxic acidic mine water was directed 
through a plastic-covered buried bed of limestone, it was 
discharged in an alkaline condition. After exposure to the 
atmosphere metal contaminants precipitated from this 
alkaline discharge much faster than they did from the 
original acid discharge. 

Since Turner and McCoy described their findings in 
1990, dozens of additional limestone treatment systems 
have been constructed ( 33 - 35 ). These passive mine water 
pretreatment systems have become known as anoxic 
limestone drains or ALD’s. In an ALD, mine water is 
made to flow through a bed of limestone gravel that has 
been buried to limit inputs of atmospheric oxygen. The 
containment caused by the burial also traps CO z within the 
treatment system, allowing the development of high C0 2 
partial pressures (36). 

Water quality data from an ALD in western Penn¬ 
sylvania are shown in table 5 and figure 7. This ALD is a 
rectangular bed of limestone gravel that is 37 m long by 
6 m wide by 1 m deep. The limestone bed is covered with 
filter fabric and 1 m of clay. No organic matter was 
incorporated into the limestone system. Water samples 
were collected from the ALD influent and effluent and at 
four locations within the ALD. The influent mine water 
contained high concentrations of ferrous iron and Mn and 
a small amount of alkalinity. As the mine water flowed 
through the ALD, pH and concentrations of calcium and 
alkalinity increased while other measured parameters were 
unchanged. Between the influent and effluent locations, 
changes in concentrations of alkalinity (137 mg-L' 1 ) and 
Ca (58 mg’L' 1 ) were in stoichiometric agreement with 
those expected from CaC0 3 dissolution. 


Table 5.—Chemistry of mine water flowing through the Howe 
Bridge anoxic limestone drain, January 23, 1992 


Parameter 

In 

Well 1 

Well 2 

Well 3 

Well 4 

Eff 

PH. 

5.9 

6.1 

6.4 

6.5 

6.5 

6.3 

Alkalinity . .. 

39 

75 

141 

179 

183 

176 

Ca. 

140 

150 

183 

201 

206 

198 

Fe 2+ . 

249 

237 

246 

246 

245 

244 

Fe 3+ . 

<1 

<1 

<1 

<1 

<1 

<1 

Mn. 

34 

33 

34 

34 

34 

34 

Al. 

<1 

<1 

<1 

<1 

<1 

<1 

Mg. 

90 

87 

91 

91 

90 

90 

Na. 

11 

11 

11 

11 

11 

11 

S0 4 . 

1175 

1175 

1200 

1150 

1200 

1200 

C0 2 . 

6.3 

4.0 

4.7 

4.3 

4.7 

NA 


NA Not available. 


NOTE.—Water flows linearly from the influent (In) through wells 
1, 2, 3, and 4 and out the effluent (Eff). C0 2 values are the partial 
pressure percentages (atmosphere) of gas samples collected from 
the headspace within the sampling wells. No gas sample could 
be collected for the effluent because it is an open pipe. 



Figure 7.—Concentrations of Ca, and alkalinity for water as It 
flows through the Howe Bridge ALO. Water flows linearly from 
influent to effluent 


Dissolution of CaC0 3 within the ALD was greater than 
would be expected from an open system in equilibrium 
with atmospheric concentrations of C0 2 (0.035%). An 
equilibrated open system would only produce alkalinity in 
the range of 50 to 60 mg»L' 1 , and increase Ca concen¬ 
trations by 4 to 8 mg»L _1 . Observations of elevated C0 2 
gas concentrations within the ALD, and the higher sol¬ 
ubility of CaC0 3 within the ALD indicate that the ALD 
acts as a closed system. 

Concentrations of alkalinity and Ca changed little be¬ 
tween the third well and the ALD effluent. This obser¬ 
vation suggests that water within the ALD was already in 
equilibrium with CaC0 3 by the time it reached the third 
well location. Thus, the amount of alkalinity that can be 
generated by this ALD is limited to a maximum value that 
is a function of the C0 2 partial pressures within the ALD. 
Similar observations of solubility-limited alkalinity gen¬ 
eration by an ALD have also been made at a second site 
in western Pennsylvania (36). 

Sulfate Reduction 

When mine water flows through an anaerobic envi¬ 
ronment that contains an organic substrate, the water 
chemistry can be affected by bacterial sulfate reduction. 
In this process, bacteria oxidize organic compounds using 
sulfate as the terminal electron sink and release hydrogen 
sulfide and bicarbonate, 

2CH 2 0 + S0 4 2 " - H 2 S + 2HC0 3 (N) 

where CH z O is used to represent organic matter. Bac¬ 
terial sulfate reduction is limited to certain environmental 

















12 


conditions (37). The bacteria require the presence of sul¬ 
fate, suitable concentrations of low-molecular weight car¬ 
bon compounds, pH >4, and the absence of oxidizing 
agents such as O^, Fe 3 * and Mn 4 \ These conditions are 
commonly satisfied in treatment systems that receive coal 
mine drainage and contain organic matter. High concen¬ 
trations of sulfate (>200 mg«L _1 ) are characteristic of 
contaminated coal mine drainage. The oxygen demand of 
organic substrates causes the development of anoxic con¬ 
ditions and an absence of oxidized forms of Fe or Mn. 
The low-molecular weight compounds that sulfate-reducing 
bacteria utilize (lactate, acetate) are common end products 
of microbial fermentation processes in anoxic environ¬ 
ments. The pH requirements can be satisfied by alkalinity 
generated by microbial activity and carbonate dissolution. 

Bacterial sulfate reduction directly affects concentra¬ 
tions of dissolved metals by precipitating them as metal 
sulfide solids. 

M 2 * + H 2 S + 2HC0 3 ~ MS + 2H 2 0 + 2C0 2 (O) 
For Fe, the formation of pyrite is also possible 

Fe 2 * + H 2 S + S° - FeS 2 + 2H* (P) 

The removal of dissolved metals as sulfide compounds 
depends on pH, the solubility product of the specific metal 
sulfide, and the concentrations of the reactants. The sol¬ 
ubilities of various metal sulfides are shown in table 6. 
Laboratory studies have verified that metal removal from 
mine water subjected to inflows of hydrogen sulfide occurs 
in an order consistent with the solubility products shown 
in table 6 (39). The first metal sulfide that forms is CuS 
followed by PbS, ZnS, and CdS. FeS is one of the last 
metal sulfides to form. MnS is the most soluble metal 
sulfide shown and is expected to form only when the con¬ 
centrations of all other metals in the table are very 
low («1 mg*L' 1 ). 

For coal mine drainage, where metal contamination is 
generally limited to Fe, Mn, and Al, the hydrogen sulfide 
produced by bacterial sulfate reduction primarily affects 


dissolved iron concentrations. Aluminum does not form 
any sulfide compounds in wetland environments and the 
relatively high solubility of MnS makes its formation 
unlikely. 


Table 6.—Solubility products of some metal sulfides 


Metal sulfide 

CdS. 

CuS. 

FeS. 

MnS. 

NiS . 

PbS. 

ZnS. 

^ee reference 38. 


Solubility product 1 

1.4 x 10' 23 
4.0 x 10 38 
1.0 x 10‘ 19 
5.6 x 10 w 
3.0 x 10* 

1.0 x 10 29 

4.5 x 10 24 


The precipitation of metal sulfides in an organic sub¬ 
strate improves water quality by decreasing the mineral 
acidity without causing a parallel increase in proton acidity. 
Proton-releasing aspects of the H 2 S dissociation process 
(HjS -» 2H* + S 2 ‘) are neutralized by an equal release of 
bicarbonate during sulfate reduction. An organic substrate 
in which 100% of the H 2 S produced by sulfate reduction 
precipitated as FeS would have no effect on the mine 
water pH or alkalinity (although acidity would decrease). 
In fact, however, the chemistry of pore water in wetlands 
constructed with an organic substrate characteristically 
has pH 6 to 8 and is highly alkaline (40-41). These alka¬ 
line conditions result, in part, from reactions involving 
hydrogen sulfide that result in the net generation of bicar¬ 
bonate. Hydrogen sulfide is a very reactive compound that 
can undergo a variety of reactions in a constructed wet¬ 
land. In most wetlands (constructed and natural), surface 
waters are aerobic while the underlying pore waters in 
contact with organic substrate are anaerobic. When sul- 
fidic pore waters diffuse from the organic substrate into 
zones that contain dissolved ferric iron, dissolved oxygen, 
or precipitated Fe and Mn oxides, the hydrogen sulfide can 
be oxidized (table 7). These reactions affect the mineral 
acidity and the alkalinity in various manners. 


Table 7.—Sinks for HjS In constructed wetlands and their net effect on mine 

water acidity and alkalinity 


Reaction 

Acidity 1 

Effect 

Alkalinity 2 

H 2 S ♦ 2 HCO 3 - - H 2 S(g) ♦ 2HC0 3 ‘ 

0 

♦ 100 

H 2 S ♦ 2HC0 3 ‘ ♦ Fe 2 * - FeS ♦ 2H 2 0 ♦ 2C0 2 

-100 

0 

H 2 S ♦ 2HCCV ♦ 2Fe 3 * - S° ♦ 2 Fe 2 * ♦ 2H 2 0 ♦ 2C0 2 

-100 

0 

H 2 S * 2HC0 3 ' * 2 Fe(OHJ 3 - S° * 2 Fe 2 * ♦ 2H 2 0 ♦ 40H' ♦ 2HC0 3 ' 

♦200 

♦300 

H 2 $ ♦ 2HC0 3 - + V*Oj - S° ♦ H 2 0 ♦ 2HCO,~ 

0 

♦ 100 

H 2 S ♦ 2HC0 3 " ♦ FeS ♦ 'A0 2 - FeS 2 ♦ H 2 0 ♦ 2HC0 3 ' 

0 

♦100 

H 2 S ♦ 2HC0 3 " ♦ 20 2 - S0 4 2 * ♦ 2H 2 0 ♦ 2C0 2 

0 

0 


1 Effect based on change in mineral acidity. 

2 Effect based on summed change in bicarbonate and hydroxyl alkalinity. 














13 


Table 8 shows the chemistry of surface water and sub¬ 
strate pore water samples collected from a wetland con¬ 
structed with limestone and spent mushroom compost. 
Spent mushroom compost consists of a mixture of spoiled 
hay, horse manure, corn cobs, wood chips, and limestone. 
At the wetland used in this example, 10 to 15 cm of lime¬ 
stone sand was covered with 20 to 50 cm of compost and 
planted with cattails. Water flowed through the wetland 
primarily by surface paths; no efforts were made to force 
the water through the compost. This design is typical 
of many compost wetlands constructed in northern 
Appalachia during the last 10 years. The data shown in 
table 8 were collected 15 months after the wetland was 
constructed. 


Table 8.—Surface and pore water chemistry 
at the Latrobe wetland 


Parameter 

Pore water 1 

Surface water 2 

Mean 

Std dev 

Mean 

Std dev 

Al. 

1 

5 

35 

5 

Ca. 

467 

188 

308 

29 

Fe 2+ . 

215 

183 

73 

39 

Fe 3+ . 

2 

9 

24 

16 

h 2 s. 

37 

75 

<1 

0 

Mg. 

175 

48 

166 

9 

Mn. 

24 

10 

42 

2 

Na. 

11 

10 

5 

1 

S0 4 . 

1,674 

532 

1,967 

115 

Acidity 3 . 

493 

340 

503 

86 

Alkalinity . 

885 

296 

0 

0 

Net Alkalinity 4 ... 

392 

NAp 

-503 

NAp 

pH. 

6.8 

.8 

3.1 

.1 


NAp Not applicable. 

Std dev Standard deviation. 

1 A total of 52 water samples were collected on July 25 and 
August 11, 1988, by the dialysis tube method. Metals were ana¬ 
lyzed for every sample. Field pH was measured for 29 samples. 
Alkalinity was measured for nine samples. 

2 Six samples collected in July and August 1988. 

3 Calculated from pH, Fe 2+ , Fe 3+ , Al, Mn, and HjS for pore 
water samples and measured by the Hj 0 2 method for surface 
water samples. 

4 Average alkalinity minus average acidity. The nine pore 
water samples for which alkalinity was measured had a mean net 
alkalinity of 653 mg/L (std dev * 590). 

Surface water at the study site had low pH and high 
concentrations of Fe, Al, and Mn (table 8). Compared 
with the surface water, the substrate pore water had higher 
pH, higher concentrations of alkalinity, ferrous iron, 
calcium, and hydrogen sulfide, and lower concentrations of 
sulfate, ferric iron, and aluminum. On average, the pore 
water had a net alkalinity while the surface water had a 
net acidity. The alkalinity cf the pore water appeared to 
result from a combination of limestone dissolution and 
sulfate reduction. The average alkalinity calculated to 
result from these processes was 703 mg'L' 1 , a value that 


corresponded reasonably well with the measured difference 
in acidity, 895 mg«L _1 . 6 

Compared with surface water, substrate pore water 
contained elevated concentrations of ferrous iron. High 
concentrations of Fe 2 * likely resulted from the dissolution 
of ferric oxyhydroxides at the redox boundary. FeOOH 
can be reduced by direct heterotrophic bacterial activity 
(< 2 ), 

CH 2 0 + 4FeOOH + H 2 0 -♦ 4Fe 2+ + 80H“ + C0 2 (Q) 
and also by H 2 S that results from sulfate reduction. 

H 2 S + 2FeOOH -* 2Fe 2+ + 40H~ + S° (R) 

In both cases, the solubilization of ferric hydroxides results 
in the release of OH", which acts to raise pH to cir- 
cumneutral levels and also reacts with dissolved C0 2 to 
form bicarbonate. Reduction of ferric hydroxide has no 
effect on the net acidity of the mine water because the 
increase in alkalinity is exactly matched by an increase in 
mineral acidity. If the Fe-enriched pore water diffuses 
into an aerobic zone, the ferrous iron content should 
oxidize, hydrolyze, and reprecipitate as ferric oxyhydroxide. 

4Fe 2+ + 80H’ + 0 2 - 4FeOOH + 2H 2 0 (S) 

Because the pore water has circumneutral pH and is 
strongly buffered by bicarbonate, the removal of iron by 
oxidation processes from pore water as it diffuses into 
aerobic surface waters should occur rapidly. Indeed, 
during the summer months, when the data in table 8 were 
collected, comparisons of the wetland influent and effluent 
indicated that the wetland decreased both concentrations 
of iron and total acidity on every sampling day (figure 8). 
The decrease in acidity indicates that alkaline pore water 
was mixing with surface water and neutralizing acidity. 
The decrease in concentrations of Fe in the surface water 
indicates that elevated concentrations of Fe 2 * observed in 
the pore water were rapidly removed in surface water 
environments. 

ALUMINUM REACTIONS IN MINE WATER 

Aluminum has only one oxidation state in aquatic 
systems, +3. Oxidation and reduction processes, which 
complicate Fe and Mn chemistry, do not directly affect 

6 The difference between surface and pore water concentrations of 
sulfate averaged 293 mg*L _1 , which is equivalent to 305 mg^L' 1 
CaCOj alkalinity (reaction N); the difference in calcium concentrations 
averaged 159 mg*L -1 , which is equivalent to 398 mg»L _1 CaC0 3 
alkalinity (reaction M). 




















TOT 


14 


-Oi 




Figure 8.—Influent and effluent concentrations at the Latrobe wetland during the summer of 1988. A, Fe; B, acidity. 


concentrations of dissolved Al. Instead, concentrations of 
A1 in mine waters are primarily influenced by the solubility 
of Al(OH)j (23, 43). At pH levels between 5 and 8, 
Al(OH) 3 is highly insoluble and concentrations of dissolved 
A3 are usually <1 mg'L' 1 . At pH values <4, Al(OH) 3 is 
highly soluble and concentrations >2 mg*!/ 1 are possible. 

The passage of mine water through highly oxidized 
or highly reduced environments has no effect on 


concentrations of Al unless the pH also changes. In those 
cases where the pH of mine water decreases (due to iron 
oxidation and hydrolysis), concentrations of Al can in¬ 
crease because of the dissolution of alumino-silicate clays 
by the acidic water. When acidic mine water passes 
through anaerobic environments, the increased pH that 
can result from carbonate dissolution or microbial activity 
causes the precipitation of Al(OH) 3 . 


CHAPTER 3. REMOVAL OF CONTAMINANTS BY PASSIVE TREATMENT SYSTEMS 


Chapter 2 described chemical and biological processes 
that decrease concentrations of mine water contaminants 
in aquatic environments. The successful utilization of 
these processes in a mine water treatment system depends, 
however, on their kinetics. Chemical treatment systems 
function by creating chemical environments where metal 
removal processes are very rapid. The rates of chemical 
and biological processes that underlie passive systems are 
often slower than their chemical system counterparts and 
thus require that mine water be retained longer before it 
can be discharged. Retention time is gained by building 
large systems such as wetlands. Because the land area 
available for wetlands on minesites is often limited, the 
sizing of passive treatment systems is a crucial aspect of 
their design. Unfortunately, in the past, most passive 
treatment systems have been sized based on guidelines 
that ignored water chemistry or on available space, rather 
than on comparisons of contaminant production by the 
mine water discharge and expected contaminant removal 
by the treatment system. Given the absence of quantita¬ 
tive sizing standards, wetlands have been constructed that 
are both vastly undersized and oversized. 

In this chapter, rates of contaminated removal are 
described for 13 passive treatment systems in western 
Pennsylvania. The systems were selected to represent the 
wide diversity of mine water chemical compositions that 
exist in the eastern United States. The rates that are 
reported from these sites are the basis of treatment system 
sizing criteria suggested in chapter 4. 


The analytical approach used to quantify the perform¬ 
ance of passive treatment systems in this chapter differs 
from the approach used by other researchers in several 
respects. First, contaminant removal is evaluated from a 
rate perspective, not a concentration perspective. Second, 
changes in contaminant concentrations are partitioned into 
two components: because of dilution from inputs of fresh¬ 
water, and because of chemical and biological processes in 
the wetland. In the evaluations of wetland performance, 
only the chemical and biological components are consid¬ 
ered. Third, treatment systems, or portions of systems, 
were included in the case studies only if contaminant 
concentrations were high enough to ensure that contam¬ 
inant removal rates were not limited by the absence of the 
contaminant. These unique aspects of the research are 
discussed in further detail below. 

EVALUATION OF TREATMENT SYSTEM 
PERFORMANCE 

To make reliable evaluations of wetland performance, 
a measure should be used that allows comparison of con¬ 
taminant removal between systems that vary in size and 
the chemical composition and flow rate of mine water they 
receive. In the past, concentration efficiency (CE%) has 
been a common measure of performance (11-12). Using 
iron concentration as an example, the calculation is 




15 


CE% « x 100 (2) 

^ e in 

where the subscripts "in" and "efP represent wetland in¬ 
fluent and effluent sampling stations and Fe concentra¬ 
tions are in milligram per liter. 

Except in carefully controlled environments, CE% is a 
very poor measure of wetland performance. The efficiency 
calculation results in the same measure of performance 
for a system that lowers Fe concentrations from 300 to 
100 mg»L' x as one that lowers concentrations from 3 to 
1 mg»L _1 . Neither the flow rate of the drainage nor the 
size of the treatment system are incorporated into the cal¬ 
culation. As a result, the performances of systems have 
been compared without accounting for differences in flow 
rate (which vary from <10 to >1000 Lamin' 1 ) or for dif¬ 
ferences in system size (which vary from <0.1 to >10 ha) 
( 12 ). 

A more appropriate method for measuring the per¬ 
formance of treatment systems calculates contaminant 
removal from a loading perspective. The daily load of 
contaminant received by a wetland is calculated from the 
product of concentration and flow rate data. For Fe, the 
calculation is 

Fe (g • d -1 ) in = 1.44 x flow (L • min -1 ) 

x Fe (mg • L-') in ) (3) 

where g* d' 1 is gram per day and 1.44 is the unit conver¬ 
sion factor needed to convert minutes to days and milli¬ 
grams to grams. 

The contaminant load is apportioned to the down flow 
treatment system by dividing by a measure of the system’s 
size. In this study, treatment systems are sized based on 
their surface area (SA) measured in square meter, 


Fe (g • m -2 • d _1 ) in = Fe (g • d'^/SA. (4) 

The daily mass of Fe removed by the wetland between two 
sampling stations, Fe(g* d' 1 )^, is calculated by comparing 
contaminant loadings at the two points, 

Fe fe * - (Fc g • - (Fe g • d-V (5) 

An area-adjusted daily Fe removal rate is then calculated 
by dividing the load removed by the surface area of the 
treatment system lying between the sampling points, 

Fe (g • m~ 2 • d -1 ) rem = Fe (g • d'^/SA. (6) 

To illustrate the use of contaminant loading and con¬ 
taminant removal calculations, consider the hypothetical 
water quality data presented in table 9. 

In systems A and B, changes in Fe concentrations are 
the same (60 mg^L' 1 ), but because system B receives four 
times more flow and thus higher Fe loading, it actually 
removes four times more Fe from the water. The concen¬ 
tration efficiencies of the two wetlands are equivalent, but 
the masses of Fe removed are quite different. 

Data are shown for system C for three sampling dates 
on which flow rates and influent iron concentrations vary. 

On the first date (Cl), the wetland removes all of the Fe 
that it receives. On the next two dates (C2 and C3), Fe 
loadings are higher and the wetland effluent contains Fe. 
From an efficiency standpoint, performance is best on the 
first date and is worst on the third date. From an Fe- 
removal perspective, the system is removing the least 
amount of Fe on the first date. On the second and third 
dates, the wetland removes similar amounts of iron (2,880 
and 3,024 g*d~ J ). Variation in effluent chemistry results, 
not from changes in wetland’s Fe-removal performance, 
but from variation in influent Fe loading. 


Table 9.—Hypothetical wetland data and performance evaluations 




Wetland 


Fe Concentration 

Fe Loading 

Fe removal 


System 

size, 

Flow rate 

In 

Eff 

In 

Eff 

performance 



m 2 

L*min' 1 

mg'L' 1 

mg'L' 1 

Kg*d' 1 

Kg-d 1 

CE 

Rate 









% 

g*m' 2 *d' 1 

A . . , 


400 

10 

100 

40 

1.4 

0.6 

60 

2.2 

B . . 


400 

40 

100 

40 

5.8 

2.3 

60 

8.6 

Cl .. 


500 

30 

40 

<1 

1.7 

<0.1 

99 

3.5 

C 2 . . 


500 

80 

35 

10 

4.0 

1.2 

71 

5.8 

C3 . . 


500 

150 

30 

16 

6.5 

3.5 

47 

6.0 

n .. 


750 

50 

100 

25 

7.2 

1.8 

75 

7.2 


In Influent. 

Eft Effluent. 

CE Concentration efficiency. 















16 


Lastly, consider a comparison of wetland systems of dif¬ 
ferent sizes. System D removes more iron than any wet¬ 
land considered (5,400 g'd' 1 ), but it is also larger. One 
would expect that, all other factors being equal, the largest 
wetland would remove the most Fe. When wetland area 
is incorporated into the measure by calculating area- 
adjusted Fe removal rates (gram per square meter per 
day), System B emerges as the most efficient wetland 
considered. 

DILUTION ADJUSTMENTS 

Contaminant concentrations decrease as water flows 
through treatment systems because chemical and biolog¬ 
ical processes remove contaminants from solution and 
because the concentrations are diluted by inputs of fresh¬ 
water. To recognize and quantify the removal of contam¬ 
inants by biological and chemical processes in passive 
treatment systems, it is necessary to remove the effects 
of dilution. Ideally, studies of treatment systems include 
the development of detailed hydrologic and chemical 
budgets so that dilution effects are readily apparent. In 
practice, the hydrologic information needed to develop 
these budgets is rarely available, except when systems 
are built for research purposes. Treatment systems con¬ 
structed by mining companies and reclamation groups are 
rarely designed to facilitate flow measurements at all water 
sampling locations, so estimating dilution from hydrologic 
information^k-highly inaccurate or impossible. 

An alternative method for distinguishing the effects 
of dilution from those of chemical and biological processes 
is through the use of a conservative ion (44-45). By de¬ 
finition, the concentration of a conservative ion changes 
between two sampling points only because of dilution or 
evaporation. Changes in concentrations of contaminant 
ions that proportionately exceed those of conservative ions 
can then be attributed to biological and chemical wetland 
processes. 

In this study, Mg was used as a conservative ion. Mag¬ 
nesium was considered a good indicator of dilution in 
these systems for both theoretical and empirical reasons. 


In northern Appalachia, concentrations of Mg in coal mine 
drainage are often >50 mg*L' 2 , while concentrations in 
r ainfall are <1 mg*L" 1 and in surface runoff are usually 
<5 mg^L' 1 . Magnesium is unlikely to precipitate in pas¬ 
sive treatment systems because the potential solid pre¬ 
cipitates, MgS0 4 , MgCOj, and CaMgtCOj)* do not form 
at the concentrations and pH conditions found in the 
systems (25). While biological and soil processes exist that 
may remove Mg in wetlands, their significance is negligi¬ 
ble relative to the high Mg loadings that most mine water 
treatment systems in northern Appalachia receive. The 
average Mg loading for wetland systems included in this 
study was ~7,000 g Mg*m* 2 *yr -1 . The uptake of dis¬ 
solved Mg by plants in constructed wetlands can only 
account for 5 to 10 g Mg»m' a# yr _1 . This estimate as¬ 
sumes that the net primary productivity of the constructed 
wetlands is 2,000 g»m' 2 *yr' 1 dry weight (46) and that the 
Mg content of this biomass is 0.25% to 0.50% (47). The 
estimate ignores mineralization processes that would 
decrease the net retention of Mg to lower values. Most 
constructed wetlands have a clay base that can adsorb Mg 
by cation exchange processes, but the total removal of 
Mg by this process is limited to about 100 g* m' 2 . This 
estimate assumes that the mine water is in contact with a 
5-cm-deep day substrate that has a density of 13 g* cm' 3 , 
a cation exchange capacity of 25 meq per 100 g, and 50% 
of the available sites are occupied by Mg (48). These con¬ 
servative calculations indicate that less than 2% of the 
annual Mg loading at the study sites is likely affected by 
biological and soil processes within the systems. 

Empirical data also indicate that Mg is conservative in 
the wetlands monitored in this study. Table 10 shows 
influent and effluent concentrations of major noncontam¬ 
inant ions at eight constructed wetlands. No precipita¬ 
tion had occurred in the study area for 2 weeks previous 
to collection of the samples, so dilution from rainfall, 
surface water, or shallow ground water seeps was minimal. 
Magnesium was the most conservative ion measured. 
Concentrations of Mg changed by <5% with flow through 
every wetland, while concentrations of all other ions mon¬ 
itored changed by at least 15% at at least one site. 


Table 10.—Influent and effluent concentrations of Ca, Mg, Na, and sulfate at eight constructed wetlands 


_Ca_ _Mg_ _Na_ _SOj_ 

In, Eff, Change, In, Eff, Change, In, Eff, Change, In, Eff, Change, 
mg'L' 1 mg-L* 1 % mg*L' 1 mg*L 4 % mg-L' 1 mg-l' 1 % mg*L* 1 mg*L‘ x % 


Donegal. 244 241 -1 81 79 -2 6 6 0 729 729 0 

Emlenton_ 429 433 *-1 308 306 -1 11 10 -2 2,810 2,770 -1 

FH. 122 189 +55 51 51 0 5 7 +2 1,125 842 -25 

Gourley. 117 120 +3 114 117 +3 3 4 +6 1,000 1,030 +3 

Latrobe . 244 256 +14 127 125 -2 6 11 +8 1,525 1,225 -20 

PineyA. 416 426 +2 251 262 +4 15 16 +4 2,190 2,120 -3 

PineyB. 355 354 0 217 216 0 27 27 -2 2,050 2,100 +2 

Somerset .... 307 469 +53 _ 312 312 _0_6_ 7 +15 2,740 2,300 -16 

Eff Effluent. 

In Influent. 


FH Friendship Hill National Historical Site. 















17 


Changes in concentrations of Mg were used to adjust 
for dilution effects by the following method. For each set 
of water samples from a constructed wetland, a dilution 
factor (DF) was calculated from changes in concentrations 
of Mg between the influent and effluent station: 

DF = Mg,, /Mg. . (7) 

Contaminant concentrations were adjusted to account for 
dilution using the DF. When only an influent flow rate 
was available, the chemical composition of the effluent 
water sample was adjusted. For Fe, the adjustment cal¬ 
culation was 

&Fc da =■ Fe m - (Fe,„/DF) (8) 

where AFe DA is expressed in milligram per liter. When 
only an effluent flow rate was available, the chemical com¬ 
position of the influent water sample was adjusted, 

AFe DA = (Fein xDF ) - Fe eff ( 9 ) 

Because most of the DF values were <1.00, the adjust¬ 
ment procedures generally resulted in smaller estimates of 
changes in contaminant concentrations than would have 
been calculated without the dilution adjustment. 

Rates of contaminant removal, expressed as gram per 
square meter per day, were then calculated from the 
dilution-adjusted change in concentrations, the flow rate 
measurement liter per minute, and the SA of the system, 
in square meter 

Fe(g • m -2 • day' 1 )^ = (AFe DA xFlow 

x 1.44 )/SA. (10) 

LOADING LIMITATIONS 

A primary purpose of this chapter is to define the 
contaminant removal capabilities of passive treatment 


systems. Accurate assessments of these capabilities re¬ 
quire that the treatment systems studied contain excessive 
concentrations of the contaminants. A system that is com¬ 
pletely effective (lowers a contaminant to <2 mg-L 1 ) 
may provide an indication that contaminant removal occurs 
(if dilution is not the cause of concentration changes), but 
cannot provide an estimate of the capabilities of the re¬ 
moval processes, as the rate of contaminant removal may 
be limited simply by the contaminant loading rate. For 
example, in table 9, the removal rate of Fe for wetland Cl 
is 3.5 g»m' 2 »d' 1 . This rate is not an accurate estimate 
of the capability of the wetland to remove Fe because 
the loading rate on this day was also only 3.5 g*m' a «d‘ l . 
The data from Cl are not sufficient to estimate whether 
the wetland could have removed 10 or 100 g»m' 2 »d _1 of 
Fe. Only when the wetland is overloaded with Fe (days 
C2 and C3), can the Fe removal capabilities of the wetland 
be assessed. 

The Morrison passive treatment system demonstrates 
the necessity of recognizing both dilution and loading- 
limiting situations in the evaluation of the kinetics of metal 
removal processes. The Morrison system consists of an 
anoxic limestone drain followed by a ditch, a settling pond, 
and two wetland cells. Figure 5, previously presented in 
chapter 2, shows average concentrations of Fe, Mn, and 
Mg at the sampling stations. Iron loading and removal 
rates for the sampling stations are shown in table 11. The 
treatment system decreased concentrations of Fe from 
151 mg»L~ l at the system influent station (the ALD dis¬ 
charge) to <1 mg«L _1 at the final wetland effluent sta¬ 
tion. Most of the change in Fe chemistry occurred in the 
ditch, a portion of the system that only accounted for 4% 
of the total treatment system SA. Calculations of the rate 
of Fe removal based on the entire treatment system re¬ 
sulted in a value of 13 g*m' a »d -1 . Because this removal 
rate is equivalent to the load, it does not represent a 
reliable approximation of the system’s Fe-removal capa- 
blity. Only when an Fe removal rate is calculated for the 
ditch, an area where Fe loading exceeded Fe removal, 
does an accurate assessment of the Fe removal capabilities 
result. 


Table 11.-Average concentration* of Fe, Mn, and Mg at the Morrison passive treatment system 



Cumulative 

Row, 


Concentration, 


Removal rate 1 , 

Station 

area, m 2 

L*m _1 


mg*L' 1 


g*m 

- 2 *d 1 




Fe 

Mn 

Mg 

Fe 

Mn 

Influent . 

0 

6.6 

151 

42 

102 

NA 

NA 

Ditch Effluent .... 

43 

NA 

56 

37 

91 

19.2 

0.17 

Pond Effluent .... 

461 

NA 

5 

24 

72 

2.3 

0.14 

Final Effluent .... 

1,076 

NA 

<1 

71 

71 

1.3 

0.13 


NA Not available. 

1 Removal rate based on cumulative area. 










18 


Concentrations of Mn at the Morrison effluent station 
were generally above discharge limits. Manganese was 
detectable in every effluent water sample (>.4 mg»L' 1 ) 
and >2 mg*L _1 in 75% of the samples. Thus, it was 
reasonable to evaluate the kinetics of Mn removal based 
on the SA of the entire treatment system. Concentrations 
of Mg, however, decreased with flow through the treat¬ 
ment system, suggesting an important dilution component. 
Effluent water samples contained, on average, 31% lower 
concentrations of Mg than did the influent samples. On 
several occassions when the site was sampled in conjunc¬ 
tion with a rainstorm, differences between effluent and in¬ 
fluent concentrations of Mg were larger than 50%. Meas¬ 
urements of metal removal by the Morrision treatment 
system that did not attempt to account for dilution would 
significantly overestimate the actual kinetics of metal 
removal processes. 

Dilution adjustments were possible for every set of 
water samples collected from a treatment system because 
concentrations of Mg were determined for every water 
sample. Problems with loading limitations, however, could 
not be corrected at every site. At two sites where com¬ 
plete removal of Fe occurred, the Blair and Donegal wet¬ 
lands, the designs of the systems were not conducive for 
the establishment of intermediate sampling stations. For 
these two systems, no Fe removal rates were calculated 
because complete removal of Fe occurred over an unde¬ 
termined area of treatment system. 

STUDY SITES 

The design characteristics of the 13 passive treatment 
systems monitored during this study are shown in table 12. 


At four of the sites, acidic mine water was pretreated with 
anoxic limestone drains (ALD’s) before it flowed into 
constructed wetlands. The construction materials for the 
wetlands ranged from mineral substances, such as clay and 
limestone rocks, to organic substances such as spent mush¬ 
room compost, manure, and hay bales. Cattails (Typha 
latifolia and, less commonly, T. angustifolia ) were the most 
common emergent plants growing in the systems. Three 
sites contained few emergent plants. Most of the wetland 
systems consisted of several cells or ponds connected seri¬ 
ally. Two systems, however, each consisted of a single 
long ditch. 

The mean influent flow rates of mine drainage at the 
study sites ranged from 7 to 8,600 L»min _1 (table 12). 
The highest flow rates occurred where drainage discharged 
from abandoned and flooded underground mines. The 
lowest flow rates occurred at surface mining sites. Esti¬ 
mated average retention times ranged from 8 h to more 
than 30 days. 

The average chemistry of the influents to the 16 con¬ 
structed wetlands are shown in table 13. Data from 15 
sampling points are shown. At the REM site, two dis¬ 
charges are treated by distinct ALD-wetland systems that 
eventually merge into a single flow. The combined flows 
arc referred to as REM-Lower. Mine water at the Howe 
Bridge system is characterized at two locations. The 
"upper" analysis describes mine water discharging from an 
ALD that flows into aerobic settling ponds. The Tower* 
analysis describes the chemistry of water flowing out of the 
last settling pond and into a large compost-limestone 
wetland that is constructed so that mine water flows in a 
subsurface manner. 


Table ^.- Construction characteristics of the constructed wetlands 


Site 

Constructed 

year 

Design 

Substrate 

Emergent 

vegetation 

m z 

Water 

depth, 

cm 

Row 

rate , 1 

L*min ' 1 

Est. ret 
time , 2 
days 

Donegal. 

1987 

Pond, 8 Cells 

LS, SMC 

Typha 

8,100 

15 

501 

1.7 

Cedar . 

1989 

5 Cells 

Clay, LS 

.. do. 

1,360 

15 

156 

0.9 

Keystone . 

1989 

Ditch 

Topsoil 

None 

4,200 

100 

8,606 

.3 

Blair. 

1989 

Ditch 

Manure, straw 

Mixed 

1,080 

5 

11 

3.4 

Shade . 

1989 

ALD, 2 Cells 

LS 

None 

880 

10 

10 

6.4 

Piney. 

1987 

1 Cell 

HB 

Mixed 

2,500 

50 

468 

1.9 

Morrison . 

1990 

ALD, 3 Cells 

Clay, manure 

Typha 

1,075 

30 

7 

33.9 

Emlenton. 

1987 

9 Cells 

LS, manure 

.. do. 

643 

50 

55 

4.1 

Somerset. 

1984 

2 Cells 

HB, LS, SMC 

.. do. 

1,005 

15 

47 

2.2 

Howe. 

1991 

ALD, 3 Cells 

Clay, LS, SMC 

None 

3,000 

50 

130 

8.0 

Latrobe . 

1987 

3 Cells 

HB, LS, SMC 

Typha 

2,800 

15 

86 

3.4 

REM . 

1992 

2 ALDs, 9 Cells 

SMC 

.. do. 

4,849 

30 

206 

4.9 

FH. 

1988 

6 Cells 

LS. SMC 

.. do. 

667 

15 

15 

4.6 


Est. Estimated. 


FH Friendship Hill National Historical Site. 

HB Haybales. 

LS Limestone, 

ret. Retention. 

SA Surface area of wet area. 

SMC Spent mushroom compost. 

1 Average values. 

2 Calculated from the water holding capacity and influent flow rate. 


















19 


Tabto 13.—Average chemical characteristic* of Influent water at the constructed wetlands 
(sites are arranged according to the net acidity) 


Site 

Number of 
samples 

pH 

Aik 

Fe 

Composition, mg«L _1 

Mn A1 

Mg 

SO, 

Net Acidity , 1,2 
mg-L ' 1 

Donegal. 


7.1 

202 

5 

8 

<1 

81 

738 

-182 

Cedar . 


6.3 

336 

92 

2 

<1 

54 

1,251 

-140 

Keystone . 


6.3 

142 

37 

<1 

<1 

14 

330 

-73 

Blair. 


6.2 

166 

52 

30 

<1 

77 

645 

-51 

Shade . 


6.0 

31 

<2 

22 

<1 

125 

966 

-17 

Piney. 


5.8 

60 

1 

15 

<1 

225 

1.845 

-6 

Morrison . 


6.3 

271 

150 

42 

<1 

102 

1,087 

75 

REM - L. 

20 

6.1 

128 

190 

50 

<1 

118 

1,275 

258 

Howe - Lower. 

13 

5.6 

22 

185 

34 

<1 

91 

1,128 

312 

Emlenton. 


4.7 

15 

89 

77 

8 

249 

2,317 

320 

Somerset. 

.... 43 

4.4 

0 

162 

50 

3 

193 

1,691 

373 

Howe - Upper. 

. . . . 13 

6.2 

160 

272 

39 

<1 

105 

1,315 

375 

REM-Lower . 

9 

3.5 

0 

246 

92 

2 

171 

1,875 

496 

Latrobe . 

43 

3.5 

0 

125 

32 

43 

125 

1,655 

617 

REM - R. 

. . . . 18 

5.5 

57 

473 

130 

3 

232 

2,495 

867 

FH. 

.... 73 

2.6 

0 

153 

9 

58 

85 

1,733 

929 


Aik Alkalinity. 

FH Friendship Hill National Historical Site. 
’CaC 03 equivalent. 

2 Negative values indicate alkaline conditions. 


Ten of the influents to the constructed wetlands had pH 
>5 and concentrations of alkalinity >25 mg«L' 1 . The 
alkaline character of five of these discharges resulted from 
pretreatment of the mine water with ALD’s. The high 
concentrations of alkalinity contained by five discharges 
not pretreated with ALD’s arose from natural geochemical 
reactions within the mine spoil (Donegal and Blair) or the 
flooded deep mine (Cedar, Keystone, and Pincy). For 
mine waters that contained appreciable alkalinity, the 
principal contaminants were Fe and Mn. 

Concentrations of alkalinity for six of the influents 
were high enough to result in a net alkaline conditions 
(negative net acidity in table 13). A seventh alkaline 
influent, Morrison, was only slightly net acidic. For these 
seven influents, enough alkalinity existed in the mine 
waters to offset the mineral acidity associated with Fe 
oxidation and hydrolysis. 

Nine of the influents were highly acidic. Five of the 
acidic influents contained alkalinity, but mineral acidity 
associated with dissolved Fe and Mn caused the solutions 
to be highly net acidic. These inadequately buffered 
waters were contaminated with Fe and Mn. Four of the 
waters contained no appreciable alkalinity (pH <4.5) and 
high concentrations of acidity. Mine waters with low pH 
were contaminated with Fe, Mn, and AJ. 

EFFECTS OF TREATMENT SYSTEMS 
ON CONTAMINANT CONCENTRATIONS 


the systems decreased Fe concentrations by more than 
50 mg^L' 1 . The largest change in Fe occurred at the 
Howe Bridge system where concentrations decreased by 
197 mg-L' 1 . From a compliance perspective, the most 
impressive decrease in Fe occurred at the Morrison system 
where 151 mg'L' 1 decreased to <1 mg^L* 1 . 

Fourteen of the passive systems received mine water 
contaminated with Mn. Eleven of these systems decreased 
concentrations of Mn. Changes in Mn were smaller than 
changes in Fe. The largest change in Mn concentration, 
31 mg'L' 1 , occurred at the Morrison site. Only the 
Donegal treatment system discharged water that con¬ 
sistently met effluent criteria for Mn (<2 mg*L _1 ). Both 
the Shade and Blair wetland effluents flowed into settling 
ponds which discharged water in compliance with regu¬ 
latory criteria. On occassions, the discharges of the 
Morrison and Pincy treatment systems met compliance 
criteria. 

Every wetland system decreased concentrations of 
acidity. The Morrison system, which received mine water 
that contained 75 mg«L‘ l acidity, always discharged net 
alkaline water. None of the constructed wetlands that 
received highly acidic water (net acidity >100 mg*L' 1 ) 
regularly discharged water with a net alkalinity. During 
low-flow periods, the Somerset, Latrobe, and FH systems 
discharged net alkaline water. The largest change in 
acidity occurred at the Somerset wetland where concen¬ 
trations decreased by an average 304 mg-L* 1 . 


The effects of the treatment systems on contaminant 
concentrations are shown in table 14. Every system de¬ 
creased concentrations of Fe. At four sites where the 
original mine discharge contained elevated concentrations 
of Fe, the final discharges contained <1 mg^L" 1 . Nine of 


DILUTION FACTORS 

While contaminant concentrations decreased with flow 
through every constructed wetland, concentrations of Mg 
also decreased at many of the sites. Decreases in Mg 






















20 


indicated that part of the improvement in water quality 
was because of dilution. Average dilution factors for the 
treatment systems are shown in table 15. For 9 of the 17 
systems, average dilution factors were 0.95 to 1.00 and 
dilution adjustments were minor. At the remaining eight 
systems, mean DF values were less than 0.95 and dilution 
adjustments averaged more than 5%. Water quality data 
from the Morrison and Somerset constructed wetlands 
were adjusted, on average, by more than 25%. 

Dilution factors varied widely between sampling days. 
Dilution adjustments were higher for pairs of samples 
collected in conjuction with precipitation events or thaws. 
Every system was adjusted by more than 5% on at least 
one occassion (see minimum dilution factors in table 15). 
Adjustments of more than 20% occurred on at least one 
occasion at 13 of the 17 study sites. 


Few dilution adjustments were >1.00 (see maximum 
dilution factors in table 15). Of the 390 dilution factors 
that were calculated for the entire data set, 13 exceeded 
1.05. These high dilution factors could have resulted from 
evaporation or freezing out of uncontaminated water with¬ 
in the treatment system, from temporal changes in water 
chemistry, or from sampling errors. Most of the high 
dilution factors were associated with rainstorm events, sug¬ 
gesting temporal changes in water quality. When dilution 
factors were >1.00, the calculated rates of contaminant 
removal were greater than would have been estimated 
without any dilution adjustment. Because of the limited 
number of sample pairs with high dilution factors, their 
presence did not markedly affect the average contaminant 
removal rates for the constructed wetland study areas. 


Table 14.—Mean water quality (or sampling stations at the constructed wetlands 


Site 

Sampling 

station 

n 1 

pH 

Fe 

Mn 

Acidity 

Mg 

Donegal. 


6 

6.4 

34 

9 

NAp 

83 


Wetland influent 

29 

7.1 

5 

8 

NAp 

81 


Effluent 

28 

7.4 

<1 

2 

NAp 

80 

Cedar . 


26 

6.3 

92 

2 

NAp 

54 


Effluent 

27 

6.4 

41 

2 

NAp 

53 

Keystone. 


28 

6.3 

37 

1 

NAp 

14 


Effluent 

28 

6.4 

32 

1 

NAp 

14 

Blair. 


12 

6.2 

52 

30 

NAp 

77 


Effluent 

8 

7.0 

<1 

5 

NAp 

59 

Shade . 


20 

6.0 

2 

23 

NAp 

128 


LC effluent 

20 

6.8 

<1 

10 

NAp 

122 

Piney. 


21 

6.4 

32 

25 

NAp 

201 


Wetland influent 

39 

5.8 

1 

15 

NAp 

225 


Wetland effluent 

39 

6.1 

<1 

11 

NAp 

225 

Morrison . 


24 

6.3 

151 

42 

75 

102 


Ditch 

24 

6.4 

56 

37 

64 

91 


Effluent 

24 

6.6 

<1 

11 

-1 

71 

REM-L. 


20 

6.1 

190 

50 

258 

118 


Left effluent 

20 

3.8 

84 

48 

225 

112 

Emlenton. 


46 

4.7 

89 

77 

320 

249 


Effluent 

40 

3.2 

15 

73 

271 

234 

Somerset. 


43 

4.4 

162 

50 

373 

193 


Effluent 

40 

5.5 

18 

33 

69 

139 

Howe. 


13 

6.0 

265 

37 

373 

101 


Upper effluent 

13 

5.6 

185 

34 

312 

91 


Lower effluent 

13 

6.2 

68 

33 

112 

91 

REM-lower . 


9 

3.5 

246 

92 

496 

171 


Effluent 

9 

2.9 

115 

88 

436 

166 

Latrobe . 


43 

3.5 

125 

32 

617 

125 


Cell 3 effluent 

43 

3.7 

56 

29 

343 

122 

REM-R. 


18 

5.5 

473 

130 

867 

232 


Right effluent 

18 

3.3 

338 

113 

712 

201 

FH. 


73 

2.6 

153 

10 

929 

85 


Effluent 

73 

2.9 

137 

10 

674 

85 


FH Friendship Hill National Historical Site. 


LC Limestone cell. 

NAp Not applicable, 
dumber of samples. 

^e flow-weighted average of two discharges. 




















21 


Table 15.—Dilution factors for the constructed wetlands 


Site 

Average 

Sd 

Minimum 

Maximum 

Donegal. 

0.99 

0.05 

0.76 

1.04 

Cedar . 

0.99 

0.03 

0.92 

1.05 

Keystone. 

0.99 

0.04 

0.91 

1.15 

Blair. 

0.83 

0.10 

0.70 

1.01 

Shade . 

0.96 

0.08 

0.76 

1.09 

Piney. 

1.00 

0.06 

0.92 

1.31 

Morrison Ditch ., 

0.87 

0.18 

0.40 

1.05 

Morrison Wetland 

0.69 

0.25 

0.27 

1.12 

REM-L. 

0.95 

0.09 

0.70 

1.13 

Howe Lower .... 

1.00 

0.10 

0.80 

1.25 

Emlenton. 

0.94 

0.09 

0.66 

1.04 

Somerset. 

0.73 

0.30 

0.30 

1.76 

Howe Upper .... 

0.89 

0.08 

0.73 

0.99 

REM-Lower .... 

0.93 

0.09 

0.72 

1.01 

Latrobe . 

0.95 

0.08 

0.75 

1.14 

REM-R. 

0.86 

0.16 

0.36 

1.00 

FH. 

1.00 

0.12 

0.58 

1.34 


FH Friendship Hill National Historical Site. 

REMOVAL OF METALS FROM ALKALINE 
MINE WATER 

Rates of Fe and Mn removal for the study systems are 
shown in table 16. Significant removal of Fe occurred at 
every study site. Fe removal rates were directly correlated 
with pH and the presence of bicarbonate alkalinity (fig¬ 
ure 9). These two water quality parameters are closely 
related because the buffering effect of bicarbonate alka¬ 
linity causes mine waters with >50 mg»L alkalinity to 
typically have a pH between 6.0 and 6.5. Within the group 
of sites that received alkaline mine water, there was not a 


significant relationship between the Fe removal rate and 
the concentration of alkalinity. 

Removal of Fe at the alkaline mine water sites ap¬ 
peared to occur principly through the oxidation of ferrous 
iron and the precipitation of ferric hydroxide (reaction A, 
chapter 2). Mine water within the systems was turbid 
with suspended ferric hydroxides. By the cessation of the 
studies, each of the alkaline water sites had developed 
thick accumulations of iron oxyhydroxides. Laboratory 
experiments, discussed in chapter 2, demonstrated that 
abiotic ferrous iron oxidation processes are rapid in aer¬ 
ated alkaline mine waters. No evidence was found that 
microbially-mediated anaerobic Fe removal processes, 
which require the presence of an organic substrate, con¬ 
tributed significantly to Fe removal at the alkaline sites. 
Fe removal rates at the REM wetlands, which were con¬ 
structed with fertile compost substrates, did not differ 
from rates at sites constructed with mineral substrates 
(Morrison, Howe-Upper, Keystone). 

Rates of Fe removal averaged 23 g»m' 2 *d‘ l at the six 
sites that contained alkaline, Fe-contaminated water. Four 
of the alkaline systems displayed similar rates despite 
widely varying flow conditions, water chemistry and sys¬ 
tem designs. The Keystone system, a deep plantless ditch 
that lowered Fe concentrations in a very large deep mine 
discharge by 5 mg L 1 , removed Fe at a rate of 
21 g*m -2 * d' 1 . The shallow-water Morrison ditch, which 
decreased concentrations of Fe in a low-flow seep by al¬ 
most 100 mg*L~ l , had an average Fe removal rate of 
19 g«m' 3 »d" 1 . The REM-L and REM-R wetlands, which 
were constructed almost identically, but received water 
with contaminant concentrations and flow rates that var¬ 
ied by 200%, displayed Fe removal rates of 20 and 
28 g»m' 2 *d‘ l . 


Table 16.—Fe and Mn removal rate* at constructed wetland 


Site 


Fe removal rate 



Mn removal rate 


Mean 

Std dev 

n 

sig? 1 

Mean 

Std dev 

n 

sig? 

Donegal. 

NAp 

NAp 

NAp 

NAp 

0.50 

0.25 

9 

yes 

Cedar . 

6.3 

2.2 

7 

yes 

0.17 

0.41 

7 

no 

Keystone. 

20.7 

5.1 

15 

yes 

NAp 

NAp 

NAp 

NAp 

Blair. 

NAp 

NAp 

NAp 

NAp 

0.43 

0.37 

6 

yes 

Shade . 

NAp 

NAp 

NAp 

NAp 

0.72 

0.64 

17 

yes 

Piney. 

NAp 

NAp 

NAp 

NAp 

1.07 

1.34 

33 

yes 

Morrison Dit. 

19.2 

10.6 

24 

yes 

0.17 

0.41 

24 

yes 

Morrison Wet. 

NAp 

NAp 

NAp 

NAp 

0.20 

0.18 

24 

yes 

REM-L. 

28.3 

5.7 

20 

yes 

•0.05 

0.13 

20 

no 

Howe-Lower... 

8.1 

1.9 

13 

yes 

0.06 

0.16 

13 

no 

Emlenton. 

9.1 

3.3 

39 

yes 

•0.09 

0.19 

39 

no 

Somerset. 

5.0 

4.9 

34 

yes 

-0.01 

0.54 

34 

no 

Howe-Upper. 

42.7 

82 

13 

yes 

-0.43 

0.49 

13 

no 

REM-Lower . 

12.0 

3.4 

9 

yes 

-0.05 

0.14 

9 

no 

Latrobe . 

2.1 

1.0 

21 

yes 

0.03 

0.09 

21 

no 

REM-R. 

20.1 

4.0 

18 

yes 

0.10 

0.33 

18 

no 

PH •«•»•••••«•••••**••• 

0.5 

0.5 

73 

yes 

0.00 

0.02 

73 

no 


NAp Not applicable. 

FH Friendship Hill National Historical Site. 

n Sample size. 

sig? Significant at 0.05 level. 

Std dev Standard deviation. 

‘Yes, rate is significantly greater than zero (t-test); no, rate is not significantly greater than zero (t-test). 







































22 


Two alkaline mine water sites varied considerably from 
the other sites in their Fe removal capabilities. The Cedar 
Grove wetland removed Fe at a rate of 6 g»m' 2 »d _1 , 
while the Howe Bridge Upper site removed Fe at a rate 
of 43 g»m‘ 2 »d' 1 . The Cedar Grove system consists of a 
series of square cells that may have more short-circuiting 
flow paths than the rectangular-shaped cells of the other 
systems. The Cedar Grove system also contains less aera¬ 
tion structures than the other systems. Mine water at the 
site upwells from a flooded underground mine into a pond 
that dicharges into a three-cell wetland. Limited topo¬ 
graphic relief prevented the inclusion of structures that 
efficiently aerate the water (i.e., waterfalls, steps). The 
Howe Bridge Upper system, in contrast, very effectively 
aerates water. Drainage drops out of a 0.3-m-high pipe, 
flows down a cascading ditch and through a V-notch weir 
before it enters a large settling pond. Because the rate of 
abiotic ferrous iron oxidation is directly proportional to 
the concentration of dissolved oxygen, insufficient oxygen 
transfer may explain the low rate of Fe removal at the 
Cedar Site, while exceptionally good oxygen transfer at the 
Howe Bridge Upper site may explain its high rate of Fe 
removal. 



INFLUENT ALKALINITY 


At sites where the buffering capacity of bicarbonate 
alkalinity exceeded the mineral acidity associated with iron 
hydrolysis, precipitation of Fe did not result in decreased 
pH. This neutralization was evident at the Morrison, 
Cedar, Keystone, Blair, Piney, and Donegal sites (ta¬ 
ble 14). At the Howe Bridge and REM wetlands, the 
mine water was insufficiently buffered and iron hydrolysis 
eventually exhausted the alkalinity and pH fell to low 
levels. The effluents of both REM systems had pH <3.5. 
The Howe Bridge Upper system discharged marginally 
alkaline water (<25 mg*L' 1 alkalinity; pH 5.6). Spot 
checks of the pH of surface water 20 m into the Howe 
Bridge Lower wetland (which receives the Upper system 
effluent) always indicated pH values <3.5. 

Significant removal of Mn only occurred at five of the 
constructed wetlands (table 13). Each of these sites re¬ 
ceived alkaline mine water (Figure 10). Each site also 
either received water with low concentrations of Fe (Piney 
and Shade) or developed low concentrations of Fe within 
the treatment system (Blair, Donegal, and Morrison). 


> 

o 

2 

UJ 

oc 




-50 0 50 100 150 200 250 300 350 400 

INFLUENT ALKALINITY 


Figure 9.—Relationship between mean Fe removal rates and 
A, mean influent pH and B, mean Influent alkalinity concen¬ 
trations. Vertical bars are one standard error above and below 
the mean. "H-L* Is the Howe-Lower site. 


Figure 10.—Relationship between mean Mn removal rates and 
A mean influent pH and B, mean influent alkalinity concen¬ 
trations. Vertical bars are one standard error above and below 
mean. Fe values next to the bars are effluent Fe 2 * values. 


















23 


Alkaline sites that contained high concentrations of Fe 
throughout the treatment system (Howc-Upper, REM-L, 
REM-R, and Cedar), did not remove significant amounts 
of Mn. The Morrison ditch, which contained water with 
an average 56 mg^L' 1 Fc, had a significant Mn removal 
rate. This rate, however, was derived from an average 
dilution-adjusted decrease in Mn concentrations of only 
1.2 mg»L _1 or 3% of the influent concentrations. Because 
of uncertainities with sampling, analysis, and dilution- 
adjustment procedures that could reasonably bias Mn data 
by 2-3%, the authors do not currently place much practical 
confidence in this value. 

The five sites that markedly decreased concentrations 
of Mn had variable designs. The Donegal wetland has a 
thick organic and limestone substrate and is densely veg¬ 
etated with cattails. The Blair and Morrison wetlands 
contain manure substrates and are densely vegetated with 
emergent vegetation. The Piney wetland was not con¬ 
structed with an organic substrate and includes deep open 
water areas and shallow vegetated areas. The Shade treat¬ 
ment system contains limestone rocks, no organic sub¬ 
strate, and few emergent plants. Thus, chemical aspects 
of the water, not particular design parameters, appear to 
principally control Mn removal in constructed wetlands. 

The removal of Mn from aerobic mine waters appeared 
to result from oxidation and hydrolysis processes. Black 
Mn-rich sediments were visually abundant in the Shade, 
Donegal, and Blair wetlands. As discussed in chapter 2, 
the specific mechanism by which these oxidized Mn solids 
form is unclear. The amorphous nature of the solids pre¬ 
vented identification by standard X-ray diffraction meth¬ 
ods. However, samples of Mn-rich solids collected from 
the Shade and Blair wetlands were readily dissolved by 
alkaline ferrous iron solutions, indicating the presence of 
oxidized Mn compounds. 

Mn 2 * can reportedly be removed from water by its 
sorption to charged FeOOH (ferric oxydroxide) particles 
(23, 30). If this process is occurring at the study wetlands, 
it is not a significant sink for Mn removal. The bottoms 
of the Morrison ditch, Howe-Upper, Cedar, REM-L, and 
REM-R wetlands were covered with precipitated FeOOH 
and the mine water within these wetlands commonly con¬ 
tained 5 to 10 mg'L' 1 of suspended FeOOH (difference 
of the Fe content of unfiltered and filtered water samples). 
After mine water concentrations were adjusted to reflect 
dilution, no removal of Mn was indicated at four of the 
sites and very minor removal of Mn occurred at the fifth 
site (Morrison ditch). 

Although the processes that remove Mn and Fc from 
alkaline mine water appears to be mechanistically similar 
(both involve oxidation and hydrolysis reactions), the ob¬ 
served kinetics of the metal removal processes arc quite 
different. In the alkaline mine waters studied, Mn removal 
rates were 20 to 40 times slower than Fe removal. 


The presence or absence of emergent plants in the wet¬ 
lands did not have a significant effect on rates of either Fe 
or Mn removal at the alkaline mine water sites. In gen¬ 
eral, bioaccumulation of metals in plant biomass is an 
insignificant component of Fe and Mn removal in con¬ 
structed wetlands (49). The ability of emergent plants to 
oxygenate sediments and the water column (50) has been 
proposed as an important indirect plant function in wet¬ 
lands constructed to treat polluted water (57). Either 
oxygenation of the water column is not a rate limiting 
aspect of metal oxidation at the constructed wetlands that 
received alkaline mine water, or physical oxygen transfer 
processes are more rapid than plant-induced processes. 

REMOVAL OF METALS AND ACIDITY 
FROM ACID MINE DRAINAGE 

Metal removal was slower at constructed wetlands that 
received acidic mine water than at those that received 
alkaline mine water. Removal of Mn did not occur at any 
site that received highly acidic water (figure 10). Removal 
of Fc occurred at every wetland that received acidic mine 
water, but the Fe removal rates were less than one-half 
those determined at alkaline wetlands (figure 9). Because 
abiotic ferrous iron oxidation processes are extremely slow 
at pH values <5, virtually all the Fe removal observed at 
the acidic sites must arise from direct or indirect microbial 
activity. Microbially-mediated Fe removal under acidic 
conditions is, however slower than abiotic Fe-removal 
processes under alkaline conditions. 

Wetlands that treat acidic mine water must both pre¬ 
cipitate metal contaminants and neutralize acidity. At 
most wetland sites, acidity neutralization was the slower 
process. At the Emlenton and REM wetlands, Fe removal 
processes were accompanied on every sampling occasion 
by an increase in proton acidity which markedly decreased 
pH (see figure 4 A, chapter 2). Mine water pH occasion¬ 
ally decreased with flow through the Latrobe and Somerset 
wetlands. Thus, for the wetlands included in this study, 
the limiting aspect of acid mine water treatment was the 
generation of alkalinity or the removal of acidity (which 
were considered in this report to be equivalent, sec chap¬ 
ter 2). The best measure of the effectiveness of the acid 
water treatment systems was through the calculation of 
acidity removal rates. 

Acidity can be neutralized in wetlands through the 
alkalinity-producing processes of carbonate dissolution and 
bacterial sulfate reduction. As was discussed in chapter 2, 
the presence of an organic substrate where reduced Eh 
conditions develop promotes both alkalinity-generating 
processes. In highly reduced environments where dis¬ 
solved oxygen and ferric iron are not present, carbonate 
surfaces are not passivated by FeOOH armoring. Decom¬ 
position of the organic substrate can result in elevated 


24 


partial pressures of COj and promote carbonate disso¬ 
lution. The presence of organic matter also promotes the 
activity of sulfate-reducing bacteria. 

The rates of alkalinity generated from these two 
processes in the constructed wetlands were determined 
based on dilution-adjusted changes in the concentrations 
of dissolved Ca and sulfate, the stoichiometry of the 
alkalinity-generating reactions, and measured flow rates. 
The calculations are based on the assumption that Ca con¬ 
centrations only increase because of carbonate dissolution 
and that sulfate concentrations only decrease because of 
bacterial sulfate reduction. One possible error in this 
approach is that sulfate can co-prccipitatc with ferric 
hydroxides in low-pH aerobic environments (52). The Fe 
and sulfate content of surface deposits collected from the 
constructed wetlands indicate that sulfate is incorporated 
into the precipitates collected from acidic environments 
at an average Fe:S0 4 ratio of 9.7 (table 17). If all of 
the Fe removed from mine water is assumed to precipitate 
as ferric hydroxide with a Fe:S0 4 ratio of 9.7:1, then 
changes in sulfate concentrations attributable to the co- 
predpitation process amount to only 5 to 30 mg*L‘‘ at 
the add mine water sites. Dilution-adjusted changes in 
sulfate concentrations at the Somerset, Latrobe, Friendship 
Hill (FH), and Howe-Lower wetlands were commonly 200 
to 500 mg*L'\ 

Rates of addity removal, sulfate removal and caldum 
addition for six constructed wetlands that received acidic 
mine water are shown in table 18. Significant removal of 
acidity occurred at all sites. The lowest rates of addity 
removal occurred at the Emlenton wetland. This site con¬ 
sists of cattails growing in a manure and limestone sub¬ 
strate. No sulfate reduction was indicated (the rate was 
not significantly >0). Dissolution of the limestone was 
indicated, but the rate was the lowest observed. 


Table 17.—Fe and S0 4 content of ferric oxyhydroxlde deposits; 
sites are arranged by pH 


Site pH Composition, ppm dry weight 

_Fe_ S0 4 Fe:SQ 4 

Emlenton. 3X) 471,779 64,213 7 X” 

Latrobe . 3.5 288,939 27,991 10.3 

Somerset. 3.5 461,583 48,263 9.6 

Cedar . 6.4 362,300 8,946 40.5 

Keystone. 6.6 398,337 6,888 57.8 


1 Field pH measured where substrate sample collected. 

The Latrobe, Somerset, FH, Howe-Lower, and REM 
systems were each constructed with a spent mushroom 
compost and limestone substrate. Spent mushroom com¬ 
post is a good substrate for microbial growth and has a 
high limestone content (10% dry weight). At these five 
wetlands, sulfate reduction and limestone dissolution both 
occurred at significant rates (table 18). The summed 
amount of alkalinity generated by sulfate reduction and 
limestone dissolution processes (Reactions M and N, 
chapter 2) correlated strongly with the measured rate of 
acidity removal at these four sites (r >0.90 at each site). 
At the FH wetland, 94% of the measured acidity removal 
could be explained by these two processes (figure 11). 

On average, sulfate reduction and limestone dissolution 
contributed equally to alkalinity generation at these five 
sites (51% versus 49%, respectively). The average sulfate 
removal rate calculated for the compost sites, 5.2 g 
S0 4 ' 2 »ra' 2 *d' 1 , is equivalent to a sulfate reduction rate 
of ~180 nmol*cm -3 * d' 1 . This value is consistent with 
measurements of sulfate reduction made at the constructed 
wetlands using isotope methods (41) as well as measure¬ 
ments of sulfate reduction made for coastal ecosystems 
(55). 


Table 18.—Average rates of acidity removal, sulfate removal, and calcium addition at sites receiving acidic mine water 


Site n Acidity removal rate Sulfate removal rate Calcium addition rate 


_ mean Std dev sig? 1 mean Std dev sig? mean Std dev sig? 

Emlenton. 25 3.1 2.4 yes 1.5 5.7 no 0.8 1.21 yes 

Somerset. 34 9.9 8.6 yes 5.1 5.7 yes 1.7 1.20 yes 

Howe Lower. 13 15.4 4.1 yes 8.9 7.2 yes 3.9 1.40 yes 

REM-Lower . 9 7.1 7.2 yes 2.9 2.4 yes 2.6 1.03 yes 

Latrobe . 21 6.9 4.4 yes 5.9 6.4 yes 0.9 0.07 yes 

FH. 72 7.0 3.8 yes 3.4 2.6 yes 1.2 0.80 yes 


FH Friendship Hill National Historical Site, 

n Sample size. 

Std dev Standard deviation. 

’Yes, rate is significantly greater than zero (t-test); no, rate is not significantly greater than zero (t-test). 

























25 


The highest rates of acidity removal, sulfate reduction, 
and limestone dissolution all occurred at the Howe-Lower 
site. This system differs from the others by its subsurface 
flow system. Drainage pipes, buried in the limestone that 
underlies the compost, cause the mine water to flow 
directly through the substrate. At the Somerset, Latrobe, 
REM, and FH systems, water flows surflcially through the 
wetlands. Mixing of the acidic surface water and alkaline 
substrate waters presumably occurs by diffusion processes 
at the surface-flow sites. By directly contacting contam¬ 
inated water and alkaline substrate, the Howe-Lower site 
is extracting alkalinity from the substrate at a significantly 
higher rate than occurs in surface flow systems. How long 
the Howe-Upper system can continue to generate alka¬ 
linity at the present rates is unknown. Monitoring of 
the system, currently in its third year of operation, is 
continuing. 



Figure 11 .—Measured rates of alkalinity generation and acidh 
removal at the Friendship Hill wetland. Units are g«m~ 2 *d 
CaC0 3 equivalent 


CHAPTER 4. DESIGN AND SIZING OF PASSIVE TREATMENT SYSTEMS 


Three principal types of passive technologies currently 
exist for the treatment of coal mine drainage: aerobic 
wetland systems, wetlands that contain an organic sub¬ 
strate, and anoxic limestone drains. In aerobic wetland 
systems, oxidation reactions occur and metals precipitate 
primarily as oxides and hydroxides. Most aerobic wetlands 
contain cattails growing in a clay or spoil substrate. How¬ 
ever, plantlcss systems have also been constructed and at 
least in the case of alkaline influent water, function sim¬ 
ilarly to those containing plants (chapter 3). 

Wetlands that contain an organic substrate are similar 
to aerobic wetlands in form, but also contain a thick layer 
of organic substrate. This substrate promotes chemical 
and microbial processes that generate alkalinity and neu¬ 
tralize acidic components of mine drainage. The term 
"compost wetland" is often used in this report to describe 
any constructed wetland that contains an organic substrate 
in which biological alkalinity-generating processes occur. 
Typical substrates used in these wetlands include spent 
mushroom compost, Sphagnum peat, haybales, and 
manure. 

The ALD is a buried bed of limestone that is intended 
to add alkalinity to the mine water (75, 33-34). The lime¬ 
stone and mine water are kept anoxic so that dissolution 
can occur without armoring of limestone by ferric oxy- 
hydroxides. ALD’s are only intended to generate alka¬ 
linity, and must be followed by an aerobic system in which 
metals are removed through oxidation and hydrolysis 
reactions. 

Each of the three passive technologies is most ap¬ 
propriate for a particular type of mine water problem. 
Often, they are most effectively used in combination with 


each other. In this chapter, a model is presented that is 
useful in deciding whether a mine water problem is suited 
to passive treatment, and also, in designing effective pas¬ 
sive treatment systems. 

Two sets of sizing criteria are provided (table 19). The 
"abandoned mined land (AML) criteria" are intended for 
groups that are attempting to cost-effectively decrease 
contaminant concentrations. In many AML situations, the 
goal is to improve water quality, not consistently achieve 
a specific effluent concentration. The AML sizing criteria 
are based on measurements of contaminant removal by 
existing constructed wetlands (chapter 3). Most of the 
removal rates were measured for treatment systems (or 
parts of treatment systems) that did not consistently lower 
concentrations of contaminants to compliance with OSM 
effluent standards. In particular, the Fe sizing factor for 
alkaline mine water (20 g^rn'^d' 1 ) is based on data 
from sue sites, only one of which lowers Fe concentrations 
to compliance. 

Table 19.—Recommended sizing for passive treatment systems 


AML criteria, Compliance criteria, 

Alkaline Acid Alkaline Acid 


Fe. 20 NAp 10 NAp 

Mn. 1.0 NAp 0.5 NAp 

Acidity .._NAp_7_NAp_^5 


NAp Not applicable. 

It is possible that Fe removal rates are a function of Fe 
concentration; i.e., as concentrations get lower, the size of 










26 


system necessary to remove a unit of Fe contamination 
(e.g., 1 g»d'*) gets larger. To account for this possibility, 
a more conservative sizing value for systems where the 
effluent must meet regulatory guidelines was provided 
(table 1). These are referred to as "compliance criteria." 
The sizing value for Fe, 10 g*nr 2 »d*\ is in agreement 
with the findings of Stark (77) for a constructed compost 
wetland in Ohio that receives marginally acidic water. 
This rate is larger, by a factor of 2, than the Fe removal 
rate reported by Brodie (18) for aerobic systems in 
southern Appalachia that are regularly in compliance. 

The Mn removal rate used for compliance, 
0.5 g*nr 2 »d‘ l , is based on the performance of five 
treatment systems, three of which consistently lower Mn 
concentrations to compliance levels. A higher removal 
value, 1 g»m~ 2 »d _l , is suggested for AML sites. Because 
the toxic effects of Mn at moderate concentrations 
(<50 mg«L _1 ) are generally not significant, except in very 
soft water (54), and the size of wetland necessary to treat 
Mn-contaminated water is so large, AML sites with Fe 
problems should receive a higher priority than those with 
only Mn problems. 

The acidity removal rate presented for compost wet¬ 
lands is influenced by seasonal variations that cannot 
currently be corrected with wetland design (55). This is 
not a problem for mildly acidic water, where the wetland 
can be sized in accordance with winter performance, nor 
should it be a major problem in warmer climates. In 
northern Appalachia, however, no compost wetland that 
consistently transforms highly acidic water (>300 mg*L _1 
acidity) into alkaline water is known. One of the study 
sites, which receives water with an average of 600 mg'L 1 
acidity and does not need to meet a Mn standard, has 
discharged water that only required chemical treatment 
during winter months. While considerable cost savings are 
realized at the site because of the compost wetland, the 
passive system must be supported by conventional treat¬ 
ment during a portion of the year. 

Because long-term metal-removal capabilities of passive 
treatment systems are currently uncertain, current Federal 
regulations require that the capability for chemical treat¬ 
ment exist at all bonded sites. This provision is usually 
met by placing a "polishing pond" after the passive treat¬ 
ment system. The design and sizing model does not cur¬ 
rently account for such a polishing pond. 

All passive treatment systems constructed at active sites 
need not be sized according to the compliance criteria pro¬ 
vided in table 19. Sizing becomes a question of balancing 
available space and system construction costs versus in¬ 
fluent water quality and chemical treatment costs. Mine 
water can be treated passively before the water enters a 
chemical treatment system to reduce water treatment costs 
or as a potential part-time alternative to full-time chemical 
treatment. In those cases where both passive and chemical 


treatment methodologies are utilized, many operators find 
that they recoup the cost of the passive treatment system 
in less than a year by using simpler, less expensive chem¬ 
ical treatment systems and/or by decreasing the amount of 
chemicals used. 

A flow chart that summarizes the design and sizing 
model is shown in figure 12. The model uses mine drain¬ 
age chemistry to determine system design, and contam¬ 
inant loadings combined with the expected removal rates 
in table 19 to define system size. The following text de¬ 
tails the use of this flow chart and also discusses aspects 
of the model that are currently under investigation. 

CHARACTERIZATION OF MINE 
DRAINAGE DISCHARGES 

To design and construct an effluent treatment system, 
the mine water must be characterized. An accurate meas¬ 
urement of the flow rate of the mine discharge or seep is 
required. Water samples should be collected at the dis¬ 
charge or seepage point for chemical analysis. Initial 
water analyses should include pH, alkalinity, Fe, Mn, and 
hot acidity (H 2 0 2 method) measurements. If an anoxic 
limestone drain is being considered, the acidified sample 
should be analyzed for Fe 3 * and Al, and a field meas¬ 
urement of dissolved oxygen should be made. 

Both the flow rate and chemical composition of a 
discharge can vary seasonally and in response to storm 



Figure 12.—Flow chert showing chemical determinations nec¬ 
essary for the design of passive treatment systems. 






















27 


events. If the passive treatment system is expected to 
be operative during all weather conditions, then the dis¬ 
charge flow rates and water quality should be measured 
in different seasons and under representative weather 
conditions. 

CALCULATIONS OF CONTAMINANT LOADINGS 

The size of the passive treatment system depends on 
the loading rate of contaminants. Calculate contaminant 
(Fe, Mn, acidity) loads by multiplying contaminant con¬ 
centrations by the flow rate. If the concentrations are 
milligrams per liter and flow rates are liters per minute, 
the calculation is 

[Fe,Mn,Acidity] g • d -1 = flow 

x [Fe,Mn, Acidity] x 1.44 (11) 

If the concentrations are milligrams per liter and flow 
rates are gallons per minute, the calculation is 

[Fe, Mn, Acidity] g • d _1 = flow 

x [Fe,Mn, Acidity] x 5.45 (12) 

Calculate loadings for average data and for those days 
when flows and contaminant concentrations are highest. 

CLASSIFICATION OF DISCHARGES 

The design of the passive treatment system depends 
largely on whether the mine water is acidic or alkaline. 
One can classify the water by comparing concentrations of 
acidity and alkalinity. 

Net Alkaline Water : alkalinity > acidity 
Net Acidic Water : acidity > alkalinity 

The successful treatment of mine waters with net acidities 
of 0 to 100 mg*L _1 using aerobic wetlands has been 
documented in this report and elsewhere (14, 18). In 
these systems, alkalinity either enters the treatment system 
with diluting water or alkalinity is generated within the 
system by undetermined processes. Currently, there is no 
method to predict which of these marginally acidic waters 
can be treated successfully with an aerobic system only. 
For waters with a net acidity >0, the incorporation of 
alkalinity-generating features (either an ALD or a com¬ 
post wetland) is appropriate. 

PASSIVE TREATMENT OF NET ALKALINE WATER 

Net alkaline water contains enough alkalinity to buffer 
the acidity produced by metal hydrolysis reactions. The 
metal contaminants (Fe and Mn) will precipitate given 


enough time. The generation of additional alkalinity is 
unnecessary so incorporation of limestone or an organic 
substrate into the passive treatment system is also un¬ 
necessary. The goal of the treatment system is to aer¬ 
ate the water and promote metal oxidation processes. In 
many existing treatment systems where the water is net 
alkaline, the removal of Fe appears to be limited by 
dissolved 0 2 concentrations. Standard features that can 
aerate the drainage, such as waterfalls or steps, should be 
followed by quiescent areas. Aeration only provides 
enough dissolved 0 2 to oxidize about 50 mg»L _1 Fe 2 *. 
Mine drainage with higher concentrations of Fe 2 * will 
require a series of aeration structures and wetland basins. 
The wetland cells allow time for Fe oxidation and hydrol¬ 
ysis to occur and space in which the Fe floe can settle out 
of suspension. The entire system can be sized based on 
the Fe removal rates shown in table 19. For example, a 
system being designed to improve water quality on an 
AML site should be sized by the following calculation: 

Minimum wetland size (m 2 ) 

= Fe loading (g • d -1 ) / 20 (g • m' 2 • d" 1 ). (13) 

If Mn removal is desired, size the system based on the Mn 
removal rates in table 19. Removal of Fe and Mn occurs 
sequentially in passive systems. If both Fe and Mn re¬ 
moval are necessary, add the two wetland sizes together. 

A typical aerobic wetland is constructed by planting 
cattail rhizomes in soil or alkaline spoil obtained on-site. 
Some systems have been planted by simply spreading 
cattail seeds, with good plant growth attained after 2 years. 
The depth of the water in a typical aerobic system is 10 to 
50 cm. Ideally, a cell should not be of uniform depth, 
but should include shallow and deep marsh areas and a 
few deep (1 to 2 m) spots. Most readily available aquatic 
vegetation cannot tolerate water depths greater than 
50 cm. 

Often, several wetland cells are connected by flow 
through a V-notch weir, lined railroad tie steps, or down 
a ditch. Spillways should be designed to pass the maxi¬ 
mum probable flow. Spillways should consist of wide cuts 
in the dike with side slopes no steeper than 2H:1V, lined 
with nonbiodegradable erosion control fabric, and coarse 
rip rap if high flows are expected (18). Proper spillway 
design can preclude future maintenance costs because of 
erosion and/or failed dikes. If pipes are used, small 
diameter (< 30 cm) pipes should be avoided because they 
can plug with litter and FeOOH deposits. Pipes should be 
made of polyvinyl chloride (PVC). More details on the 
construction of aerobic wetland systems can be found in a 
text by Hammer (56). 

The geometry of the wetland site as well as flow con¬ 
trol and water treatment considerations may dictate the 




28 


use of multiple wetland cells. The intercell connections 
may also serve as aeration devices. If there are elevation 
differences between the cells, the interconnection should 
dissipate kinetic energy and be designed to avoid erosion 
and/or the mobilization of precipitates. 

It is recommended that the freeboard of aerobic wet¬ 
lands constructed for the removal of Fe be at least 1 m. 
Observations of sludge accumulation in existing wetlands 
suggest that a 1-m freeboard should be adequate to con¬ 
tain 20 to 25 years of FeOOH accumulation. 

The floor of the wetland cell may be sloped up to about 
3% grade. If a level cell floor is used, then the water level 
and flow are controlled by the downstream dam spillway 
and/or adjustable riser pipes. 

As discussed in chapter 3, some of the aerobic systems 
that have been constructed to treat alkaline mine water 
have little emergent plant growth. Metal removal rates in 
these plantless, aerobic systems appears to be similar to 
what is observed in aerobic systems containing plants. 
However, plants may provide values that are not reflected 
in measurements of contaminant removal rates. For ex¬ 
ample, plants can facilitate the filtration of particulates, 
prevent flow channelization and provide wildlife benefits 
that are valued by regulatory and environmental groups. 

PASSIVE TREATMENT OF NET ACiO WATER 

Treatment of acidic mine water requires the generation 
of enough alkalinity to neutralize the excess acidity. Cur¬ 
rently, there are two passive methods for generating alka¬ 
linity: construction of a compost wetland or pretreatment 
of acidic drainage by use of an ALD. In some cases, the 
combination of an ALD and a compost wetland may be 
necessary to treat the mine water. 

ALD’s produce alkalinity at a lower cost than do 
compost wetlands. However, not all water is suitable for 
pretreatment with ALD’s. The primary chemical factors 
believed to limit the utility of ALD’s are the presence of 
ferric iron (Fe 3 *), aluminum (Al) and dissolved oxygen 
(DO). When acidic water containing any Fe 3 * or Al 
contacts limestone, metal hydroxide particulates (FeOOH 
or Al(OH) 3 ) will form. No oxygen is necessary. Ferric hy¬ 
droxide can armor the limestone, limiting its further dis¬ 
solution. Whether aluminum hydroxides armor limestone 
has not been determined. The buildup of both precipitates 
within the ALD can eventually decrease the drain perme¬ 
ability and cause plugging. The presence of dissolved 
oxygen in mine water will promote the oxidation of ferrous 
iron to ferric iron within the ALD, and thus potentially 
cause armoring and plugging. While the short-term per¬ 
formance of ALD’s that receive water containing elevated 
levels of Fe 3 *, Al, or DO can be spectacular (total 
removal of the metals within the ALD) (34), the long-term 
performance of these ALD’s is questionable. 


Mine water that contains very low concentrations of 
DO, Fe 3 * and Al (all <1 mg*L* 1 ) is ideally suited for 
pretreatment with an ALD. As concentrations of these 
parameters rise above 1 mg'L' 1 , the risk that the ALD 
will fail prematurely also increases. Recently, two ALD’s 
constructed to treat mine water that contained 20 rag^L' 1 
Al became plugged after 6-8 months of operation. 

In some cases, the suitability of mine water for pre- 
treatment with an ALD can be evaluated based on the 
type of discharge and measurements of field pH. Mine 
waters that seep from spoils and flooded underground 
mines and have a field pH > 5 characteristically have con¬ 
centrations of DO, Fe 3 *, and Al that are all <1 mg»L‘ l . 
Such sites are generally good candidates for pretreatment 
with an ALD. Mine waters that discharge from open drift 
mines or have pH <5 must be analyzed for Fe 3 * and Al. 
Mine waters with pH <5 can contain dissolved Al; mine 
waters with pH <3.5 can contain Fe 3 *. In northern 
Appalachia, most mine drainages that have pH <3 contain 
high concentrations of Fe 3 * and Al. 

PRETREATMENT OF ACIDIC WATER WITH ALD 

In an ALD, alkalinity is produced when the acidic water 
contacts the limestone in an anoxic, closed environment. 
It is important to use limestone with a high CaC0 3 content 
because of its higher reactivity compared with a limestone 
with a high MgCOj or CaMg(C0 3 ) 2 content. The lime¬ 
stones used in most successful ALD’s have 80% to 95% 
CaC0 3 content. Most effective systems have used number 
3 or 4 (baseball-size) limestone. Some systems con¬ 
structed with limestone fines and small gravel have failed, 
apparently because of plugging problems. The ALD must 
be sealed so that inputs of atmospheric oxygen are min¬ 
imized and the accumulation of C0 2 within the ALD is 
maximized. This is usually accomplished by burying the 
ALD under several feet of clay. Plastic is commonly 
placed between the limestone and clay as an additional gas 
barrier. In some cases, the ALD has been completely 
wrapped in plastic before burial (35). The ALD should be 
designed so that the limestone is inundated with water at 
all times. Clay dikes within the ALD or riser pipes at the 
outflow of the ALD will help ensure inundation. 

The dimensions of existing ALD’s vary considerably. 
Most older ALD’s were constructed as long narrow drains, 
approximately 0.6 to 1.0 m wide. A longitudinal section 
and cross section of such an ALD is shown in figure 13. 
The ALD shown was constructed in October 1990, and is 
1 m wide, 46 m long and contains about 1 m depth of 
number 4 limestone. The limestone was covered with two 
layers of 5 mil plastic, which in turn was covered with 


29 



Figure 13.—Longitudinal-section and cross-section of the Morrison ALO. Wells are for sampling purposes and have no Importance 
to drain’s functioning. 


0.3 to 3 m of on-sitc clay to restore the original surface 
topography (34, 36). 

At sites where linear ALD’s are not possible, anoxic 
limestone beds have been constructed that are 10 to 20 m 
wide. These bed systems have produced alkalinity concen¬ 
trations similar to those produced by the more conven¬ 
tional drain systems. 

The mass of limestone required to neutralize a certain 
discharge for a specified period can be readily calculated 
from the mine water flow rate and assumptions about the 
ALD’s alkalinity-generating performance. Recent USBM 
research indicates that approximately 14 h of contact time 
between mine water and limestone in an ALD is necessary 
to achieve a maximum concentration of alkalinity (57). To 
achieve 14 h of contact time within an ALD, -3,000 kg of 
limestone rock is required for each liter per minute of 
mine water flow. An ALD that produces 275 mg^L' 1 of 
alkalinity (the maximum sustained concentration thus far 
observed for an ALD), dissolves ~ 1,600 kg of limestone a 
decade per each liter per minute of mine water flow. To 
construct an ALD that contains sufficient limestone to 
insure a 14-h retention time throughout a 30-yr period, the 
limestone bed should contain —7,800 kg of limestone for 


each liter per minute of flow. This is equivalent to 30 tons 
of limestone for each gallon per minute of flow. The 
calculation assumes that the ALD is constructed with 90% 
CaCOj limestone rock that has a porosity of 50%. The 
calculation also assumes that the original mine water does 
not contain ferric iron or aluminum. The presence of 
these ions would result in potential problems with armor¬ 
ing and plugging, as previously discussed. 

Because the oldest ALD’s are only 3 to 4 yr old, it is 
difficult to assess how realistic these theoretical calcu¬ 
lations are. Questions about the ability of ALD’s to main¬ 
tain unchannelized flow for a prolonged period, whether 
100% of the CaC0 3 content of the limestone can be ex¬ 
pected to dissolve, whether the ALD’s will collapse after 
significant dissolution of the limestone, and whether inputs 
of DO that are not generally detectable with standard field 
equipment (0 to 1 mg’L' 1 ) might eventually result in 
armoring of the limestone with ferric hydroxides, have not 
yet been addressed. 

The anoxic limestone drain is one component of a pas¬ 
sive treatment system. When the ALD operates ideally, its 
only effect on mine water chemistry is to raise pH to 











































30 


circumneutral levels and increase concentrations of cal¬ 
cium and alkalinity. Dissolved Fe 2 * and Mn should be 
unaffected by flow through the ALD. The ALD must be 
followed by a settling basin or wetland system in which 
metal oxidation, hydrolysis and precipitation can occur. 
The type of post-ALD treatment system depends on the 
acidity of the mine water and the amount of alkalinity 
generated by the ALD. If the ALD generates enough 
alkalinity to transform the acid mine drainage to a net 
alkaline condition, then the ALD effluent can then be 
treated with a settling basin and an aerobic wetland. If 
possible, the water should be aerated as soon as it exits 
the ALD and directed into a settling pond. An aerobic 
wetland should follow the settling pond. The total post- 
ALD system should be sized according to the criteria 
provided earlier for net alkaline mine water. At this time, 
it appears that mine waters with acidities <150 mg*L _1 
are readily treated with an ALD and aerobic wetland 
system. 

If the mine water is contaminated with only Fe 2 * and 
Mn, and the acidity exceeds 300 mg»L~ l , it is unlikely that 
an ALD constructed using current practices will dis-charge 
net alkaline water. When this partially neutralized water 
is treated aerobically, the Fe will precipitate rapidly, but 
the absence of sufficient bufferring can result in a 
discharge with low pH. Building a second ALD, to re¬ 
charge the mine water with additional alkalinity after it 
flows out of the aerobic system, is currently not feasible 
because of the high DO content of water flowing out of 
aerobic systems. If the treatment goal is to neutralize all 
of the acidity passively, then a compost wetland should be 
built so that additional alkalinity can be generated. Such 
a treatment system thus contains all three passive tech¬ 
nologies. The mine water flows through an ALD, into a 
settling pond and an aerobic system, and then into a com¬ 
post wetland. 

If the mine water is contaminated with ferric iron 
(Fe 3 *) or Al, higher concentrations of acidity can be 
treated with an ALD than when the water is contaminated 
with only Fe 2 * and Mn. This enhanced performance re¬ 
sults from a decrease in mineral acidity because of the 
hydrolysis and precipitation of Fe 3 * and Al within the 
ALD. These metal-removing reactions decrease the min¬ 
eral acidity of the water. ALD’s constructed to treat mine 
water contaminated with Fe 3 * and Al and having acidities 
greater than 1,000 mg'L" 1 have discharged net alkaline 
water. The long-term prognosis for these metal-retaining 
systems has been questioned (34). However, even if cal¬ 
culations of system longevity (as described above) are 
inaccurate for waters contaminated with Fe 3 * and Al, their 
treatment with an ALD may turn out to be cost-effective 
when compared with chemical alternatives (35). 


When a mine water is contaminated with Fe 2 * and Mn 
and has an acidity betweem 150 and 300 mg^L' 1 , the 
ability of an ALD to discharge net alkaline water will 
depend on the concentration of alkalinity produced by the 
limestone system. The amount of alkalinity generated by 
a properly constructed and sized ALD is dependent on 
chemical characteristics of the acid mine water. An ex¬ 
perimental method has been developed that results in 
an accurate assessment of the amount of alkalinity that 
will be generated when a particular mine water contacts a 
particular limestone (58). The method involves the anoxic 
incubation of the mine water in a container filled with 
limestone gravel. In experiments at two sites, the con¬ 
centration of alkalini ty that developed in these containers 
after 48 h correlated well with the concentrations of 
alkalinity measured in the ALD effluents at both sites. 

TREATING MINE WATER WITH COMPOST 
WETLAND 

When mine water contains DO, Fe 3 * or Al, or contains 
concentrations of acidity >300 mg»L _1 , construction of a 
compost wetland is recommended. Compost wetlands 
generate alkalinity through a combination of bacterial ac¬ 
tivity and limestone dissolution. The desired sulfate- 
reducing bacteria require a rich organic substrate in which 
anoxic conditions will develop. Limestone dissolution also 
occurs readily within this anoxic environment. A substance 
commonly used in these wetlands is spent mushroom 
compost, a substrate that is readily available in western 
Pennsylvania. However, any well-composted equivalent 
should serve as a good bacterial substrate. Spent mush¬ 
room compost has a high CaC0 3 content (about 10% dry 
weight), but mixing in more limestone may increase the 
alkalinity generated by CaC0 3 dissolution. Compost sub¬ 
strates that do not have a high CaC0 3 content should 
be supplemented with limestone. The compost depth used 
in most wetlands is 30 to 45 cm. Typically, a metric ton 
of compost will cover about 3.5 m 2 to a depth of 45 cm 
thick. This is equivalent to one ton per 3.5 yd 2 . Cattails 
or other emergent vegetation are planted in the substrate 
to stabilize it and to provide additional organic matter 
to "fuel" the sulfate reduction process. As a practical tip, 
cattail plant-rhizomes should be planted well into the 
substrate prior to flooding the wetland cell. 

Compost wetlands in which water flows on the surface 
of the compost remove acidity (e.g., generate alkalinity) 
at rates of approximately 2-12 g*m" 2 *d -1 . This range in 
performance is largely a result of seasonal variation: lower 
rates of acidity removal occur in winter than in summer 
(55). Research in progress indicates that supplementing 
the compost with limestone and incorporating system 
designs that cause most of the water to flow through the 


31 


compost (as opposed to on the surface) may result in 
higher rates of limestone dissolution and better winter 
performance. 

Compost wetlands should be sized based on the re¬ 
moval rates in table 19. For an AML site, the calculation 
is 

Minimum Wetland Size (m 2 ) = 

Acidity Loading (g • d _1 /7). (14) 

In many wetland systems, the compost cells are pre¬ 
ceded with a single aerobic pond in which Fe oxidation 
and precipitation occur. This feature is useful where the 
influent to the wetland is of circumneutral pH (either 
naturally or because of pretreatment with an ALD), and 
rapid, significant removal of Fe is expected as soon as the 
mine water is aerated. Aerobic ponds are not useful when 
the water entering the wetland system has a pH <4. At 
such low pH, Fe oxidation and precipitation reactions are 
quite slow and significant removal of Fe in the aerobic 
pond would not be expected. 

OPERATION AND MAINTENANCE 

Operational problems with passive treatment systems 
can be attributed to inadequate design, unrealistic ex¬ 
pectations, pests, inadequate construction methods, or 
natural problems. If properly designed and constructed, a 


passive treatment system can be operated with a minimum 
amount of attention and money. 

Probably the most common maintenance problem is 
dike and spillway stability. Reworking slopes, rebuilding 
spillways, and increasing freeboard can all be avoided by 
proper design and construction using existing guidelines 
for such construction. 

Pests can plague wetlands with operational problems. 
Muskrats will burrow into dikes, causing leakage and 
potentially catastrophic failure problems, and will uproot 
significant amounts of cattails and other aquatic vegetation. 
Muskrats can be discouraged by lining dike inslopes with 
chainlink fence or riprap to prevent burrowing (13). 
Beavers cause water level disruptions because of damming 
and also seriously damage vegetation. They are very dif¬ 
ficult to control once established. Small diameter pipes 
traversing wide spillways ("three-log structure") and trap¬ 
ping have had limited success in beaver control. Large 
pipes with 90° elbows on the upstream end have been used 
as discharge structures in beaver-prone areas (18). Other¬ 
wise, shallow ponds with dikes with shallow slopes toward 
wide, riprapped spillways may be the best design for a 
beaver-infested system. 

Mosquitos can be a problem where mine water is alka¬ 
line. In southern Appalachia, mosquitofish (Gambusia 
affinis) have been introduced into alkaline-water wetlands. 
Other insects, such as the armyworm, have devastated 
monocultural wetlands with their appetite for cattails (59). 
The use of a variety of plants in a system will minimize 
such problems. 


CHAPTER 5. SUMMARY AND CONCLUSIONS 


The treatment of contaminated coal mine drainage 
requires the precipitation of metal contaminants and the 
neutralization of acidity. In conventional treatment sys¬ 
tems, distinctions between these two treatment objectives 
are blurred by additions of highly basic chemicals that 
simultaneously cause the rapid precipitation of metal con¬ 
taminants and the neutralization of acidity. Passive treat¬ 
ment differs from conventional treatment by its distinction 
between these two treatment objectives. It is possible to 
passively precipitate Fe contaminants from mine water, but 
have little effect on the mine water acidity. Alternatively, 
it is possible to passively add neutralizing capacity to acidic 
mine water without decreasing metal concentrations. 

Waters that contain high concentrations of bicarbonate 
alkalinity are most amenable to treatment with constructed 
wetlands. Bicarbonate acts as a buffer that neutralizes the 
acidity produced when Fe and Mn precipitate and main¬ 
tains a pH between 5.5 and 6.5. At this circumneutral pH, 
Fe and Mn precipitation processes are more rapid than 


under acidic pH conditions. Given the ability of bi¬ 
carbonate alkalinity to positively impact both the metal 
precipitation and neutralization aspects of mine water 
treatment, it is not surprising that the most noteworthy 
applications of passive treatment have been at sites where 
the mine water was net alkaline. The most successful wet¬ 
lands constructed in western Pennsylvania in the early 
1980’s treated mine waters that contained alkalinity. All 
of the early successes of the TVA were, likewise, with 
waters that were alkaline (13). Similarly, the Simco wet¬ 
land in Ohio, which has discharged compliance water for 
several years (77), receives water containing -160 mg»L _1 
alkalinity. In this study, the two treatment systems that 
met all effluent discharge requirements (Donegal and 
Blair) both received alkaline, metal-contaminated water. 

When mine water is acidic, enough alkalinity must be 
generated by the passive treatment system to neutralize 
the acidity. The most common method used to passively 
generate alkalinity is the construction of a wetland that 


32 


contains an organic substrate in which alkalinity-generating 
microbial processes occur. If the substrate contains 
limestone, as spent mushroom compost does, then alka¬ 
linity will be generated by both calcite dissolution and 
bacterial sulfate reduction reactions. These alkalinity 
generating processes are slow relative to processes that 
remove Fe. Thus, the performance of the constructed wet¬ 
lands that receive acidic water is usually limited by the rate 
at which alkalinity is generated within the substrate. While 
wetlands can significantly improve water quality, and have 
proven to be effective at moderately acidic sites, no wet¬ 
land systems that consistently and completely transform 
highly acidic water to compliance quality are known. 
Inconsistent or partial treatment indicates undersiring. 
The authors believe this is because of a lack of awareness 
of how much larger wetlands constructed to treat acidic 
water must be than ones constructed to treat alkaline 
water. The Fe and acidity removal rates measured in this 
study indicate that the treatment of 5,000 g'd* 1 of Fe in 
alkaline water requires -250 m 2 of aerobic wetland. The 
treatment of the same Fe load in acidic water (where 
treatment requires both precipitation of the Fe and neu¬ 
tralization of the associated acidity) requires -1,300 m 2 of 
compost wetland. Thus wetlands constructed to treat 
acidic water need to be six times larger than ones con¬ 
structed to treat similarly contaminated alkaline water. 

The recent development of limestone pretreatment sys¬ 
tems, e.g., the anoxic limestone drain, is a significant ad¬ 
vancement in passive treatment technology. When suc¬ 
cessful, ALD’s can lower acidities or actually transform 
acidic water into alkaline water, and markedly decrease the 
sizing demands of the wetlands constructed to precipitate 
the metal contaminants. Because limestone is inexpensive, 
the cost of an ALD-aerobic wetland passive treatment 
system is typically much less than the compost wetland 
alternative. Thus, when the influent water is appropriate, 
ALD’s should be the preferred method for generating 
alkalinity in passive treatment systems. 

Anoxic limestone drains have also been used to increase 
the performance of existing constructed wetlands. At 
many poorly performing wetlands that receive acidic water, 
the wetland was built too small to treat an acidic, metal- 
contaminated influent, but is large enough for an alkaline, 
metal-contaminated influent. One of the study sites, the 
Morrison wetland, was undersized for the highly acidic 
water that it received. As a result, the wetland effluent 
required supplemental treatment with chemicals. Since 
construction of an ALD, and its addition of 275 mg’L* 1 
of bicarbonate alkalinity to the water, the discharge of the 
wetland has been alkaline, low in dissolved metals, and 
does not require any supplemental chemical treatment. 
Similar enhancements in wetland performance through 
the addition of ALD’s have been reported elsewhere in 
Appalachia (75, 18). 


KINETICS OF CONTAMINANT 
REMOVAL PROCESSES 

This report presents an intensive analysis of con¬ 
taminant removal kinetics in passive treatment systems. 
The rates presented are generally in agreement with those 
reported by other investigators. For example, the average 
Mn-removal rate measured in this study for alkaline, 
Fe-free waters, 0.5 g»m~ 2 «d‘ l , is consistent with rates 
reported by the TVA for aerobic wetlands in southern 
Appalachia (18) and by the Pennsylvania Department of 
Environmental Resources (DER) for constructed wetlands 
in Pennsylvania (60). The average Fe-removal rate .re¬ 
ported in this study for alkaline waters, 20 g«m" 2 «d _l , is 
only slightly greater than has been reported in other 
studies. The rates of Fe removal for aerobic wetlands 
in southern Appalachia ranged from 6 to 20g«m' 2 *d' 1 
(18). Some of the lower rates reported by TVA investi¬ 
gators, however, are from wetland systems that discharge 
water with <1 mg»L' 1 Fe and thus are loading limited 
with respect to Fe. Such sites were intentionally avoided 
in this study. Stark (77), in their studies of a constructed 
wetland in Ohio, reported Fe removal rates over a range 
of loading conditions. When the wetland system dis¬ 
charged >15 mg*L* 1 Fe, and thus was overloaded with 
Fe, the removal rate averaged 21 g^m’^d" 1 . When the 
wetland effluent contained <15 mg»L _1 Fe, the removal 
rate averaged only 11 g»m' 2 *d' 1 . 

LONG-TERM PERFORMANCE 

Passive treatment systems cannot be expected to per¬ 
form indefinitely. In the long term, wetland systems will 
fill up with metal precipitates or the conditions that 
facilitate contaminant removal may be compromised. 
None of the treatment systems considered in this study 
demonstrated any downward trends in contaminant re¬ 
moval performance. Therefore, estimates of the long¬ 
term performance of passive systems must be made by 
extrapolating available data. Like the design and siring 
of passive treatment systems, estimates of long-term per¬ 
formance vary with the chemistry of the mine water. Sys¬ 
tems receiving alkaline water precipitate Fe and Mn con¬ 
taminants by oxidative processes. The rapid removal of 
Fe that occurs in alkaline treatment systems means that 
such systems will inevitably fill up. Stark (61) reports that 
the Fe sludge in a constructed wetland in Ohio is in¬ 
creasing by 3 to 4 cm per year. Similar measurements at 
Pennsylvania wetlands indicate an increase in sludge depth 
of 2 to 3 cm per year (62). These measurements suggest 
that dikes that provide 1 m of freeboard should provide 
sufficient volume for 25 to 50 years of performance. 

At some surface mines, water quality tends to improve 
within a decade after regrading and reclamation arc 


33 


completed (63-64). At these surface minesites, 25 to 
50 years of passive treatment may be adequate to mitigate 
the contaminant problem. At surface mine sites where 
contaminant production is continual, or at systems con¬ 
structed to treat drainage from underground mines or coal 
refuse disposal areas, the system can either be built with 
greater freeboard or rebuilt when it eventually fills up. 
Site conditions will determine whether it is more econom¬ 
ical to simply bury the wetland system in place and con¬ 
struct a new one, or to excavate and haul away the ac¬ 
cumulated solids for proper disposal. Disposal of these 
excavated sludges is not difficult or unduly expensive 
because the material is not considered a hazardous waste. 

Wetlands that receive acidic water, and function 
through the alkalinity-generating processes associated with 
an organic substrate, may decline in performance as the 
components of the organic substrate that generate alka¬ 
linity are exhausted. The compost wetlands described in 
this report neutralize acidity through the dissolution of 
limestone and the bacterial reduction of sulfate. Lime¬ 
stone dissolution is limited by the amount of limestone 
present in the substrate. The limestone content of spent 
mushroom compost is ~30 kg»m~* (65). If a wetland 
containing a 40 cm depth of compost generates CaC0 3 - 
derived alkalinity at a mean rate of 3 g»m' 2 »d~ l (the 
average rate measured in this study), then the limestone 
content of the compost will be exhausted in 11 years. The 
same volume of compost contains ~40 kg of organic car¬ 
bon. If bacterial sulfate reduction mineralizes 100% of 
this carbon to bicarbonate at a rate of 5 g»m' 2 «d~ l , then 
the carbon will be exhausted in 91 years. This estimate is 
increased by the carbon input of the net primary produc¬ 
tion of the wetland system, but decreased by the fact that 
some of the carbon is mineralized by reactions other than 
sulfate reduction. Studies of a salt marsh on Cape Cod, 
MA, indicated that 75% of the carbon was eventually min¬ 
eralized by sulfate reduction processes (66). Another sig¬ 
nificant factor that decreases the available carbon is that 
a portion of the carbon pool is recalcitrant. 

A realistic scenario for the long-term performance of 
a compost wetland is that sulfate reduction is linked, in 
a dependent manner, to limestone dissolution. Sulfate- 
reducing bacteria are inactive at pH less than 5 (37). 
Their activity in a wetland receiving lower pH water may 
depend, in part, on the presence of pH-buffcring supplied 
by limestone dissolution. Thus, limestone dissolution may 
create alkaline zones in which sulfate reduction can 
proceed and produce further alkalinity. If this scenario is 
accurate, then the long-term performance of a compost 
wetland may be limited by the amount of limestone in the 
substrate, or according to the above calculations, about 
11 years of performance. Under these conditions ii would 
be advisable to increase the chemical buffering capability 
of the wetland substrate by adding additional limestone 


during wetland construction. In fact, this procedure is 
commonly practiced at many constructed compost wetland 
sites. 

The performance of anoxic limestone drains has many 
aspects that make long-term expectations uncertain. An¬ 
oxic limestone drains function through the dissolution, 
and thus removal, of limestone. Eventually, this chemical 
reaction will exhaust the limestone. Long-term scenarios 
about ALD performance fail to consider the hydrologic 
implications of the gradual structural failure of the sys¬ 
tems. In large ALD’s, most of the limestone dissolution 
occurs in the upgradient portion of the limestone bed. It 
is unknown whether this preferential dissolution will 
produce partial failure of the integrity of the system or 
whether the permeability will be adversely affected. 
Another aspect that affects long-term ALD performance 
is the fact that ALD’s retain ferric iron and aluminum (34- 
35). This retention has raised concerns about the ar¬ 
moring of limestone or the plugging of flow paths long 
before the limestone is exhausted by dissolution reactions 
(34). No methods are currently available to predict exactly 
how the retention of these metals affects the performance 
of ALD’s. 

CONTINUALLY EVOLVING PASSIVE 
TECHNOLOGIES 

This document reports the current state of passive mine 
water treatment technologies. The design and sizing rec¬ 
ommendations presented herein represent current meth¬ 
odologies that will subsequently be replaced with more 
efficient techniques. For example, important experiments 
arc underway in Pennsylvania, Virginia, and West Virginia 
testing "hybrid" ALD-compost wetland systems. In these 
experimental systems, organic substrates are used to re¬ 
duce ferric iron to ferrous iron and strip dissolved oxygen 
from the water so that the mine water is suitable for flow 
through an anoxic limestone drain. If these systems prove 
successful, it may be possible to treat highly acidic water 
by cycling it between anoxic alkalinity-generating environ¬ 
ments and aerobic, metal-removal environments. Experi¬ 
mental systems using this design have recently been con¬ 
structed in western Pennsylvania (67). 

While the specific tools of passive treatment are likely 
to evolve in the coming years, the fundamental mech¬ 
anisms of passive treatment that have been identified in 
this report will probably not change markedly. Research 
has shown that the treatment of contaminated coal mine 
drainage by constructed wetlands can be explained by well- 
known chemical and biological processes. Passive treat¬ 
ment, like active treatment with chemicals, requires that 
the metal contaminants be precipitated and that the acidity 
associated with these ions be neutralized. By recognizing 
that these treatment goals need not be accomplished 


34 


<3 

simultaneously, one can focus on optimization of the 
individual objectives. As a result, the performance and 
cost effectiveness of passive treatment systems is rapidly 
improving. Today, most mine operators who install prop¬ 
erly designed passive treatment systems rapidly recoup the 


cost of their investment through decreased water treatment 
costs. There is no reason to doubt that this technology 
will continue to improve and that, over time, passive 
treatment will be used in applications that are not possible 
today. 


REFERENCES 


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1989, pp. 16I-16N. 

2. Kleinmann, R. L. P., J. R. Jones, and P. M. Erickson. An 
Assessment of the Coal Mine Drainage Problem. Paper in the Proc¬ 
eedings of the 10th Annual Conference of the Association of Abandoned 
Mine Land Programs, PA Bur. AML Reclam., 1988, pp. 1-9. 

3. Herlihy, A. T., P. R. Kaufman, M. E. Mitch, and D. Brown. 
Regional Estimates of Acid Mine Drainage Impact on Streams in the 
Mid-Atlantic and Southeastern United States. Water, Air, and Soil 
Pollution, v. 50, 1990, pp. 91-107. 

4. Huntsman, B. E, J. G. Solch, and M. D. Porter. Utilization of 
Sphagnum Species Dominated Bog for Coal Acid Mine Drainage 
Abatement. Geological Society of America (91st Annual Meeting) 
Abstracts, Toronto, Ontario, Canada, 1978, p. 322. 

5. Wieder, R. K, and G. E Lang. Modification of Acid Mine 
Drainage in a Freshwater Wetland. Paper in Proceedings of the Sym¬ 
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INT.BU.OF MINES,PGH.,PA 29916 








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